Role of organic carbon, nitrate and ferrous iron on the partitioning between denitrification and DNRA in constructed stormwater urban wetlands

Science of the Total Environment 666 (2019) 608–617

Contents lists available at ScienceDirect

Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Role of organic carbon, nitrate and ferrous iron on the partitioning between denitrification and DNRA in constructed stormwater urban wetlands Md. Moklesur Rahman ⁎, Keryn L. Roberts, Michael R. Grace, Adam J. Kessler, Perran L.M. Cook Water Studies Centre, School of Chemistry, Monash University, Clayton, Australia

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• Natural organic carbon (NOC) and ferrous iron showed significant effect on the partitioning between DNF and DNRA. • Nitrate reduction was stimulated more by the addition of NOC than acetate. • NOC enhanced DNRA more than denitrification. • Fe2+ had a significant control on the balance between denitrification and DNRA. • There was a link between DNRA and Fe2 + oxidation in Fe2+ amended slurries.

a r t i c l e

i n f o

Article history: Received 25 October 2018 Received in revised form 14 February 2019 Accepted 14 February 2019 Available online 15 February 2019 Editor: Ashantha Goonetilleke Keywords: Denitrification DNRA Wetland Organic carbon Ferrous Fe2+

a b s t r a c t Denitrification (DNF) and dissimilatory nitrate reduction to ammonium (DNRA) are two competing nitrate reduction pathways that remove or recycle nitrogen, respectively. However, factors controlling the partitioning between these two pathways are manifold and our understanding of these factors is critical for the management of N loads in constructed wetlands. An important factor that controls DNRA in an aquatic ecosystem is the electron donor, commonly organic carbon (OC) or alternatively ferrous iron and sulfide. In this study, we investigated the role of natural organic carbon (NOC) and acetate at different OC/NO− 3 ratios and ferrous iron on the partitioning between DNF and DNRA using the 15N-tracer method in slurries from four constructed stormwater urban wetlands in Melbourne, Australia. The carbon and nitrate experiments revealed that DNF dominated at all OC/NO− 3 ratios. The higher DNF and DNRA rates observed after the addition of NOC indicates that nitrate reduction was enhanced more by NOC than acetate. Moreover, addition of NOC in slurries stimulated DNRA more than DNF. Interestingly, slurries amended with Fe2+ showed that Fe2+ had significant control on the balance between DNF and DNRA. From two out of four wetlands, a significant increase in DNRA rates (p b .05) at the cost of DNF in the presence of available Fe2+ suggests DNRA is coupled to Fe2+ oxidation. Rates of DNRA increased 1.5–3.5 times in the Fe2+ treatment compared to the control. Overall, our study provides direct evidence that DNRA is linked to Fe2+ oxidation in some wetland sediments and highlights the role of Fe2+ in controlling the partitioning between removal (DNF) and recycling (DNRA) of bioavailable N in stormwater urban constructed wetlands. In our study we also measured anammox and found that it was always b0.05% of total nitrate reduction in these sediments. © 2019 Elsevier B.V. All rights reserved.

⁎ Corresponding author. E-mail addresses: [email protected], [email protected] (M.M. Rahman), [email protected] (K.L. Roberts), [email protected] (M.R. Grace), [email protected] (A.J. Kessler), [email protected] (P.L.M. Cook). https://doi.org/10.1016/j.scitotenv.2019.02.225 0048-9697/© 2019 Elsevier B.V. All rights reserved.

M.M. Rahman et al. / Science of the Total Environment 666 (2019) 608–617

1. Introduction Nitrogen (N) availability is a key factor determining primary productivity in aquatic ecosystems (Brin et al., 2015; Hardison et al., 2015). Therefore, a better understanding of processes removing and recycling N and associated factors controlling N dynamics is important for strategies to safeguard water quality as well as the health of aquatic ecosystems (Gardner and McCarthy, 2009; Hou et al., 2012). Although different nitrate reduction processes exist in aquatic ecosystems, the relative degree to which they occur is variable and can be controlled by multiple factors such as availability of organic carbon, nitrate and ferrous iron (Akunna et al., 1993; van den Berg et al., 2016; Weber et al., 2006a). Nitrate reduction in anaerobic sediments is mainly governed by two dissimilatory processes: denitrification (DNF) and dissimilatory nitrate reduction to ammonium (DNRA) (Bonaglia et al., 2016; Burgin and Hamilton, 2007; Sgouridis et al., 2011; Tomaszek and Rokosz, 2007). The ecological significance of these two competing processes is that DNF results in a permanent loss of N from an ecosystem whereas DNRA retains N in the preferred form (NH4+) for biological uptake (Dong et al., 2011; Kelly-Gerreyn et al., 2001; Tiedje, 1988; Yin et al., 2002). Wetlands, including constructed wetlands, are highly productive ecosystems rich in organic matter and effective at removing problematic nutrients such as NO3− through DNF (DeLaune et al., 2005; White and Reddy, 2003). However, some studies have shown that NO3− can be converted to NH4+ by DNRA bacteria under conditions of high organic carbon (Bonin, 1996; Nijburg et al., 1997; Tiedje, 1988). Based on empirical observations, Tiedje et al. (1983) proposed that the fate of NO3− is regulated by the OC/NO3− ratio and that DNRA is the more efficient energy conservation process. Therefore, DNRA shuttles electrons more efficiently compared to DNF under high organic carbon and low NO3− conditions due to greater biomass yields (Strohm et al., 2007). Denitrifying bacteria are favoured by increased availability of NO3− (Nijburg et al., 1997) while Schmidt et al. (2011) reported a strong correlation between OC/NO3− and DNRA activity. However, some studies have observed that increased NO3− availability favoured anaerobic ammonium oxidation (anammox) or DNRA relative to DNF (Dong et al., 2011; Kraft et al., 2014; Rich et al., 2008) whereas other studies have found no relationship between OC/NO3− ratios and fate of NO3− (Kelso et al., 1997; Stevens et al., 1998). Alternatively, in chemoautotrophic DNF and DNRA, reductants such as highly reduced sulfur (S2−) or ferrous iron (Fe2+) are used instead of organic matter as an electron donor to reduce NO3− to N2 (Fossing et al., 1995; Reyes-Avila et al., 2004; Straub et al., 1996). In addition to chemoautotrophic denitrifiers, a number of heterotrophic denitrifiers are also able to use Fe2+ as an electron donor when organic carbon is absent. Under low oxygen and organic carbon depleted conditions, an existing denitrifier can continue removing N from the system through Fe2+ oxidation (Benz et al., 1998; Straub et al., 1996). It was inferred that the majority of the bacterial isolates which are efficiently using Fe2+ as an electron donor during nitrate reduction produce N2 as the end product (Chakraborty and Picardal, 2013; Muehe et al., 2009; Straub et al., 1996). This process is often mixotrophic, which requires an organic co-substrate such as acetate (Chakraborty and Picardal, 2013; Kappler and Straub, 2005; Muehe et al., 2009). Several microbial isolates studies have shown N2 is the final product of NOx reduction coupled to Fe2+ oxidation (Chakraborty and Picardal, 2013; Kappler and Straub, 2005; Straub et al., 1996). However, recent studies have shown that NH4+ can be the final product of NOx reduction coupled to Fe2+ oxidation (Coby et al., 2011; Robertson et al., 2016; Weber et al., 2006b). Although various phylogenetic groups have been identified to potentially contribute to nitrate reduction with Fe2+ oxidation (Laufer et al., 2016; Straub et al., 2004), only Geobacter sp. has so far been isolated to carry out DNRA coupled to Fe2+ oxidation (Finneran et al., 2002; Lovley et al., 1993). Members of the Geobacter sp. are abundant in iron-rich freshwater sediments (Weber et al., 2006b) and have

609

been found to coexist and compete with heterotrophic denitrifiers in freshwater lake sediments (Melton et al., 2012; Melton et al., 2014). It has also been shown that NO2− reduction to NH4+ can occur through abiotic mechanisms (Hansen et al., 1996a; Klueglein and Kappler, 2013; Tai and Dempsey, 2009). The reaction stoichiometries for Fe2+ driven nitrate and nitrite reduction for DNF and DNRA are as follows (Robertson et al., 2016): NO3 − þ 5 Fe2þ þ 12 H2 O→1=2 N2 ðgÞ þ 5 FeðOHÞ3 þ 9Hþ

ð1Þ

ΝΟ3 − þ 8 Fe2þ þ 21 Η2 Ο→ΝΗ4 þ þ 8 FeðΟΗÞ3 þ 14 Ηþ

ð2Þ

NO2 − þ 3 Fe2þ þ 7 H2 O→1=2 N2 ðgÞ þ 3 FeðOHÞ3 þ 5Hþ

ð3Þ

NO2 − þ 6 Fe2þ þ 16 Η2 Ο→ΝΗ4 þ þ 6 FeðΟΗÞ3 þ 10 Ηþ

ð4Þ

Although some studies investigated the role of the OC/NO3− ratio and different sources of organic carbon on the proportions of DNF and DNRA in coastal sediments and chemostat cultures (Brin et al., 2015; Hardison et al., 2015; Krishna Mohan et al., 2016; Liu et al., 2016; van den Berg et al., 2016), the effect of the OC/NO3− ratio on nitrate reduction processes in freshwater constructed wetlands remains largely unknown. Also, only a few studies have focused on the link between Fe2+ availability and dissimilatory nitrate reduction processes (Coby et al., 2011; Weber et al., 2001; Weber et al., 2006a; Weber et al., 2006b). Iron is a very common metal in wetlands with concentrations ranging from tens to hundreds of μM in sediment porewaters (Emerson et al., 1999) and has been shown to use as an electron donor for nitrate reduction in some sediments (Achtnich et al., 1995; Devlin et al., 2000; Straub et al., 1996; Weber et al., 2006a). A recent study on freshwater lake sediment showed that Fe2+ addition in slurries stimulated DNRA with simultaneous decrease in DNF in some cases (Robertson and Thamdrup, 2017). However, despite the potential of Fe2+ to reduce nitrate to either N2 or NH4+, the role of Fe2+ on dissimilatory nitrate reduction in freshwater ecosystems including freshwater wetlands is still unclear. Because stormwater wetlands receive significant loads of suspended sediments coated in reactive iron, such systems have the potential to have very high Fe2+ concentrations within the sediment porewater, potentially stimulating DNRA. Previous studies used either pure sources of organic carbon including methanol, glucose, acetate, lactate, malate, citrate and oxalate (Her and Huang, 1995; Krishna Mohan et al., 2016; Liu et al., 2016; Paul et al., 1989; Tam et al., 1992; van den Berg et al., 2016) or natural sources of organic carbon such as algal organic matter, cornstalks, woodchips and cardboard fibres (Brin et al., 2015; Greenan et al., 2006; Hardison et al., 2015) for nitrate reduction. In our study, we investigated the role of both NOC and acetate at two different OC/NO3− ratios and Fe2+ availability on the partitioning of dissimilatory nitrate reduction between DNF and DNRA from four constructed stormwater wetlands in Melbourne, Australia. In this regard, slurries of surface sediments amended with organic carbon, 15N-NO3− and Fe2+ were used to test the hypothesis that DNRA would predominate under conditions of high OC/NO3− ratio and increased Fe2+ availability. 2. Materials and methods 2.1. Sampling Four stormwater constructed urban wetlands (Huntingdale Road, Koolamara Blvd, Namatjira Reserve and Cascades on Clyde) were chosen in south-east Melbourne, Australia based on a previous study in which it was inferred that nitrate concentration, organic carbon (OC) and Fe2+ were linked to DNF and DNRA (Rahman et al., 2019). From the %OC content, it was apparent that two of the wetlands, Huntingdale Road and Koolamara Blvd were characterized by relatively high %OC (9–11%) and the other two wetlands, Namatjira Reserve and Cascades

610

M.M. Rahman et al. / Science of the Total Environment 666 (2019) 608–617

on Clyde, were characterized by relatively low %OC (4–5%). From the dissolved Fe2+ concentration in sediment porewater, it was apparent that three of the wetlands, Huntingdale Road, Koolamara Blvd and Cascades on Clyde were characterized by relatively high Fe concentrations (mean 220–240 μM) and the other wetland, Namatjira Reserve was characterized by relatively low Fe concentration (mean 170 μM) (Rahman et al., 2019). Prior to collecting sediment cores for slurry experiments, natural organic carbon (NOC) samples of Juncus sp., the dominant vegetation type in all four wetlands, were collected, cleansed with ultrapure Milli-Q water and then oven dried at 50 °C. The dried NOC was ground to a powder with a ball mill grinder (Retsch Mixer Mill MM400). Intact sediment cores for slurry experiments were collected from the inlet of all four wetlands in polyethylene cylinders (6.5 cm inner diameter; 27–29 cm height) in October and November 2015 for carbon and nitrate experiments and in November 2016 and February 2017 for Fe2+ addition experiments. Immediately on collection, cores were placed in the dark and on ice, and slurries were prepared with top 1 cm sediments in the laboratory within 2 h. 2.2. Background data Physico-chemical parameters of the surface water were recorded from all sites during each sampling occasion from December 2014 to February 2016 (Rahman et al., 2019). Water column concentrations ranged from 0.7 to 24 μM for NH4+ and 1.4 to 57.3 μM for NOx (NO3− + NO2−). Sediments from all sampling sites were characterized from April–May 2015 to January–February 2016 (Rahman et al., 2019). Total Fe content in dry sediments ranged from 0.5% in Namatjira Reserve to 4.1% in Cascades on Clyde while total nitrogen ranged from 0.1 to 1.2% with most samples below 1%. The mean organic carbon content in sediments of these wetlands ranged from 4 to 11%. Parameters measured from sediment porewater showed that Fe2+ in the porewater ranged from 6 μM in Cascades on Clyde to 750 μM in Koolamara Blvd. Dissolved NH4+ in porewater ranged from below detection (0.143 μM) to a maximum of 1000 μM whereas NOx concentration ranged from 0.1 to 7.1 μM. 2.3. Slurry incubations with organic carbon Surface sediment down to 1 cm was extruded from the cores, transferred into 12.5 mL vials (Exetainer, Labco) and mixed with unfiltered site water to prepare the slurry samples. Slurry experiments amended with different substrates of OC and different concentrations of nitrate were conducted in the laboratory at room temperature (23–28 °C) to determine the potential rates of DNF and DNRA using the modified methods of Dodla et al. (2008) and Greenan et al. (2006). Briefly, to prepare a 10% slurry sample for OC and nitrate experiments, approximately 0.8 g of homogenized wet sediment was mixed with site water to a final volume of 8 mL. Four slurries were prepared per treatment and each treatment consisted of 3 replicates. Slurries were homogenized by gentle shaking. Along with the no‑carbon addition control, two carbon treatments were prepared: an acetate treatment (40 mg of acetate as anhydrous sodium acetate (CH3COONa) was added); and a NOC treatment (40 mg of NOC). The amount of NOC and acetate added to the slurries was 5% (w/w) of the sediments, which is within the range of sediment OC content (2.4–12.8%) observed in a previous study (Rahman et al., 2019), corresponding to an additional source of 5 g/L of OC. Samples in exetainer vials from all three treatments were purged with He for 5 min to ensure anoxic conditions. Thereafter, sample vials were left overnight on a shaker table at 130 rpm in the dark. The following day, each of the three treatments was amended with either high or low nitrate concentrations. For the high nitrate concentration, 250 μL of 15 mM 15N-NO3− solution (15N atom%, 99.6%, Novachem Pty Ltd., Australia) was added to each vial of three treatments (control, acetate and NOC) giving a final nitrate concentration of 420 μmol L−1. The high concentration was chosen to simulate a high nitrate inflow event

and provide insight into how concentrations affect the partitioning of nitrate reduction pathways. For the low nitrate concentration, 100 μL of 1.5 mM 15N-NO3− solution (15N atom%, 99.6%, Novachem Pty Ltd., Australia) was added to each vial of the three treatments giving a final nitrate concentration of 20 μmol L−1 in the slurry. In a preliminary experiment, we observed that 100 μmol L−1 nitrate from slurries was completely reduced within 3 h. Therefore, a 3 h period was chosen for the slurry incubations. Slurry vials for each treatment were terminated by adding 250 μL of 50% (w/w) ZnCl2 after 0, 1, 2 and 3 h. Denitrification and DNRA were calculated from the accumulation of 15N-N2 and 15NNH4+, respectively, over time (see Denitrification and DNRA rate calculations for details). 2.4. Slurry incubations with Fe2+ and acetate For slurry experiments with Fe2+, 2.5 g of homogenized wet sediments from the top 1 cm was mixed with 47.5 mL of unfiltered site water into 60 mL gas tight serum vials (5% slurries). Vials were sealed with bromobutyl rubber stoppers and slurries were purged with helium (high purity, Air Liquid) for 15 min, followed by shaking at 130 rpm overnight to remove residual oxygen and nitrate (Robertson et al., 2016). A solution of FeSO4, prepared anoxically in a glove box, was added to vials for the Fe treatment to a final concentration of 2 mM. The pH was adjusted to within 0.2 units of the control treatment (pH 7.5, 7.4, 7.4, 6.9 for Huntingdale Road, Namatjira Reserve, Koolamara Blvd and Cascades on Clyde wetlands, respectively) by adding sterile filtered 0.5 mol L−1 HCl or NaOH. Vials were then returned to the shaker table for 24 h and sampled (0, 30 min, 1, 3, 6, 12, 24 h) for Fe2+ and pH to ensure the concentration of Fe2+ in the slurry had stabilised after the addition. At the end of the preincubation for Fe and pH stability (24 h), the headspace of each vial was flushed with helium for 2 min to remove any N2 prior to the tracer addition. 15N-NO3− tracer was added to each vial to a final concentration of 0.5 mM to determine the potential rates of DNF and DNRA. In another set containing Fe2+, acetate was added (final concentration of 0.5 mM) to assess the contribution of heterotrophic organisms, which use acetate as a co-substrate for their growth during iron fuelled nitrate reduction (Chakraborty and Picardal, 2013; Kappler and Straub, 2005; Robertson et al., 2016; Straub and Buchholz-Cleven, 1998). At each time point (0, 0.5, 1, 3, 6,12, 24 h), a 0.5 mL gas sample for N2 analysis was collected from the headspace using a 0.5 mL glass syringe (Hamilton) and hypodermic needle, while simultaneously replacing the gas volume with helium. The 15N-N2 gas samples were then transferred into 12.5 mL Exetainers (Labco) containing a 4.5 mL of helium headspace and 8 mL of helium-purged ultrapure water. For dissolved Fe2+, 1 mL of slurry sample was filtered through a 0.2 μm polyethersulfone (PES) filter (Sartorius) and was preserved with ferrozine. Another 1 mL of filtered sample was preserved with 50% (w/v) ZnCl2 and stored frozen for subsequent analysis. The headspace of each vial was flushed following each sampling time point to remove any residual 15N-N2. Total 15N-N2 accumulation over time was then calculated. At the end of the experiment, 5 mL of each slurry sample was collected for 15 NH4+ after extraction with 2 M KCl to determine sediment-bound 15 NH4+ via hypobromite conversion from 15NH4+ to 15N2 (RisgaardPetersen et al., 1995; Roberts et al., 2014). Samples were extracted with 1:1 (v/v) 2 M KCl to slurry, shaken at 120 rpm for 1 h and then centrifuged. 8 mL of filtered extractant was transferred into a 12.5 mL Exetainers (Labco) and purged with helium to eliminate background N2 followed by adding 200 μL of alkaline hypobromite to convert the 15 NH4+ to 15N2. Samples were then shaken at 130 rpm for 24 h prior to analysis to ensure complete conversion of 15NH4+ to 15N2. 2.5. Denitrification and DNRA rate calculations Denitrification was estimated measuring the excess 29N2 and 30N2 accumulated in the headspace, while DNRA was estimated from the

M.M. Rahman et al. / Science of the Total Environment 666 (2019) 608–617 15 NH4+ accumulated in the slurries (Dalsgaard et al., 2000; Nielsen, 1992; Risgaard-Petersen et al., 1995). The headspaces for both DNF and DNRA were pre-treated with 50 μL of air (Roberts et al., 2014) to increase the background concentration of 14N2 in the sample and accurately allow the calculation of excess 15N-N2 from the N2 ratios of 28/ 29 and 28/30. Rates of DNF were determined from the linear production of 29N2 and 30N2 over time (Dalsgaard et al., 2000; Nielsen, 1992) and were used to calculate the rate of DNF produced from the consumption of 15N-NO3− (D15). Rates of DNRA were calculated from the linear production of 15N-NH4+ over time. The rate of DNRA (DNRA15) or production of 15N-NH4+ produced in the slurries was determined via hypobromite conversion method as mentioned above. The method has an isotope ratio precision of ~0.01%. A series of standards were prepared in the same matrix to test the recovery of 15NH4+. Recovery for all standards was 100 ± 5%. The 15N2 produced was analysed on a Sercon 20–22 isotope ratio mass spectrometer connected to a gas chromatograph.

2.6. Statistical analysis Statistical analyses were conducted using SigmaPlot 13.0. Differences in potential rates of DNF and DNRA and their ratio from three treatments were determined through one-way analysis of variance (one-way ANOVA). Significance of the statistical test result was determined at an α = 0.05 level. 3. Results 3.1. Rates of denitrification and DNRA at low and high OC/NO3− ratios Potential rates of DNF from the control, acetate and NOC treatments at high OC/NO3− ratio ranged from 2.0 ± 0.5 to 8 ± 4 μmol L−1 h−1 (Fig. 1a). Rates of DNF from all three treatments of the four wetlands were similar, however the Koolamara Blvd wetland showed a significant (p b .05) decrease in rate from the NOC treatment compared to the control. In contrast, potential rates of DNRA from all three treatments at high OC/NO3− ratio ranged from 0.2 ± 0.1 to 1.2 ± 0.4 μmol L−1 h−1 (Fig. 1c), an order of magnitude lower than DNF rates. DNRA rates increased significantly (p b .05) in the acetate treatment at Huntingdale Road wetland, which was categorized by relatively high %OC. Unlike DNF, the potential rates of DNRA increased significantly (p b .05) in the NOC treatment compared to the control of the Koolamara Blvd wetland, which was also categorized by relatively high %OC. However, the rates from the Namatjira Reserve and the Cascades on Clyde wetlands, which were categorized by relatively low % OC, did not differ significantly from the control. The DNF:DNRA ratio from all three treatments at high OC/NO3− ratio ranged from 2.3 to 25.4 with relatively higher ratios from the Huntingdale Road wetland and lower ratios from the Namatjira Reserve wetland (Fig. 1e). The DNF:DNRA ratio for the acetate and NOC treatment of the Huntingdale Road wetland and only the NOC treatment of the Koolamara Blvd wetland were significantly (p b .05) lower compared to the control. In all other cases, there was no significant difference between the treatments and the control. Potential rates of DNF from the control, acetate and NOC treatments at low OC/NO3− ratio ranged from 13.5 ± 0.7 to 68 ± 4 μmol L−1 h−1 (Fig. 1b). Rates of DNF increased from 5.5 to 16 times in the control, 4 to 20 times in the acetate and 8 to 31 times in the NOC treatment compared to those at high OC/NO3− ratio. DNF rates increased significantly (p b .05) from the NOC treatment compared to those from the control of all but the Namatjira Reserve wetland. Rates from the acetate treatment of the Cascades on Clyde wetland increased significantly (p b .05) whereas the rate from the acetate treatment of the Huntingdale Road wetland interestingly decreased significantly (p b .05) compared to those from the control. However, the DNF rates from the acetate

611

treatment of the Namatjira Reserve and the Koolamara Blvd wetlands did not differ significantly from the control. Potential rates of DNRA from all three treatments at low OC/NO3− ratio ranged from 1.7 ± 0.5 to 18 ± 2 μmol L−1 h−1 (Fig. 1d). Rates of DNRA at low OC/NO3− ratio increased from 3 to 8 times in the control, 2.5 to 8 times in the acetate and 4.5 to 32 times in the NOC treatment compared to those at high OC/NO3− ratio. Rates of DNRA from the acetate and the NOC treatments increased compared to the control. The increase for the acetate treatment above the control was significant (p b .05) only for the Namatjira Reserve and the Koolamara Blvd wetlands; however, the NOC treatment showed significant (p b .05) increases above the control for all but the Huntingdale Road wetland. The DNF: DNRA ratio from all three treatments at low OC/NO3− ratio ranged from 3 to 25 (Fig. 1f). The DNF:DNRA ratio at low OC/NO3− ratio decreased for all three treatments in all wetlands but the Huntingdale Road wetland for which there was no difference to the control (Fig. 1f). However, this decrease in the ratio was significant (p b .05) only for the acetate and the NOC treatment of the Namatjira Reserve and the Koolamara Blvd wetlands and only for the NOC treatment of the Cascades on Clyde wetland. 3.2. Slurry experiments with Fe2+ and acetate The addition of Fe2+ to the slurries led to an increase (36 to 58% of the added amount) in dissolved Fe2+ concentration from 716 to 1200 μmol L−1 compared to the control. However, the majority of the added iron was moved from solution into mineral phases during the 24 h pre-incubation period. During this pre-incubation period, 67% of the added dissolved Fe2+ was moved from solution into mineral phases from the Fe2+ treatment of the Huntingdale Road, 75% of the Namatjira Reserve, 82% of the Koolamara Blvd and 96% of the Cascades on Clyde wetlands. Similarly, dissolved Fe2+ was also moved from solution into mineral phases from the Fe2+ + acetate treatment, corresponding to 62%, 66%, 68% and 96% for the Huntingdale Road, Namatjira Reserve, Koolamara Blvd and Cascades on Clyde wetlands, respectively. Desorption of particle-bound Fe(II) occurred from both the Fe2+ and Fe2+ + acetate treatments of the Huntingdale Road wetland (Fig. 2a, b). It is apparent from Fig. 2a and b that the concentrations of Fe2+ was gradually decreasing during the pre-incubation period (0 to ~25–30 h). However, the initial concentration during the incubation period in these two figures (~25–30 h to 60 h) was higher than the final concentration of pre-incubation period which suggests that particle-bound Fe (II) was released during this period. Denitrification rates from all four treatments of Fe2+ experiments ranged from 3.7 ± 0.5 to 47 ± 8 μmol L−1 h−1 with the lowest rate from the Fe2+ treatment; the highest rate was observed in the acetate treatment (Fig. 3a). In all wetlands, the addition of Fe2+ had a negative effect on denitrification, which is evident from the lower denitrification rate in the Fe2+ treatment compared to the control and in the Fe2+ + acetate compared to acetate treatment. In contrast, the addition of acetate to slurry samples enhanced DNF and the rate increases for the acetate treatment were significant (p b .05) compared to the control for all but the Namatjira Reserve wetland. Only the Cascades on Clyde wetland showed a significant (p b .05) increase in DNF rate for the Fe2 + + acetate treatment compared to the control. DNF rates from the Huntingdale Road wetland were lower compared to the other three wetlands while the Namatjira Reserve wetland showed relatively higher rates than the other three wetlands from all four treatments. Rates of DNRA from all four treatments of the Fe2+ experiments ranged from 0.7 ± 0.1 to 76 ± 11 μmol L−1 h−1 with the lowest rate from the control treatment whereas the highest rate was observed in the acetate treatment (Fig. 3b). Clearly, Fe2+ had a positive effect on DNRA, however, DNRA rates from the Fe2+ treatment compared to the control increased significantly (p b .05) only from the Huntingdale Road and the Cascades on Clyde wetlands. In addition to Fe2+, acetate also exhibited a positive influence on DNRA. Rates of DNRA from the

612

M.M. Rahman et al. / Science of the Total Environment 666 (2019) 608–617

80

-1

75

70 15 10

High NO3- treatments

b

Control Acetate NOC

DNF (µm ol Lslurry -1 h-1)

-1

DNF (µm ol Lslurry h )

80

Low NO3- treatments

a

*

*

Control Acetate NOC

60

*

40

*

20 *

5 * 0

0 Huntingdale

Namatjira

Koolomara

Cascades

Huntingdale

Namatjira

Sites

25

c

Control Acetate NOC

24.5 2.0 1.5 *

1.0

*

DNRA (µm ol Lslurry -1 h-1)

DNRA (µm ol Lslurry -1 h-1)

25.0

Koolamara

Cascades

Sites

d

Control Acetate NOC

20

*

15 * 10

*

*

* 5

0.5 0.0

0 Huntingdale

Namatjira

Koolamara

Cascades

Huntingdale

Namatjira

Sites 25

25

e

Control Acetate NOC

DNF:DNRA

DNF:DNRA

*

10

f

Cascades

Control Acetate NOC

20

20

15

Koolamara

Sites

15 *

*

10

* 5

*

5

*

*

* 0

0 Huntingdale

Namatjira

Koolamara

Cascades

Huntingdale

Namatjira

Koolamara

Cascades

Sites

Sites

Fig. 1. Rates of denitrification at low and high NO− 3 treatments (a, b), DNRA (c, d) and their relative proportion (e, f) from slurries amended with acetate and NOC. Asterisks (*) above bars represent significant (p b .05, ANOVA) differences between control and acetate- and NOC-amended slurries. Error bars are ±1SD (n = 3).

acetate treatment were significantly (p b .05) higher compared to the control in all except the Namatjira Reserve wetland. The combined effect of Fe2+ and acetate on DNRA was the strongest treatment effect, which led to a significant (p b .05) increase in DNRA rates compared to the control for all four wetlands. Rates of DNRA were relatively lower from the Cascades on Clyde wetland compared to the other three wetlands whereas the Huntingdale Road wetland showed higher DNRA than the other three wetlands. The DNF:DNRA ratio from Fe2+ experiments ranged from 0.1 to 12 (Fig. 3c). The ratios for the Fe2+ treatment of all wetlands and for the acetate treatment of all except the Namatjira Reserve wetland were significantly lower than the control. However, the ratios were significantly (p b .05) lower than the control for the Fe2+ + acetate treatment only for the Huntingdale Road and the Namatjira Reserve wetlands. Ratios were relatively lower from the Huntingdale Road wetland compared to the other three wetlands due to higher DNRA rates whereas the Cascades on Clyde wetland showed higher ratios than the other three wetlands mainly due to lower DNRA

rates. DNF:DNRA ratios were invariably higher from the control compared to other three treatments. 4. Discussion 4.1. Effect of OC/NO3− ratio on the partitioning between denitrification and DNRA Potential rates of DNF observed in this study are comparable to those observed from slurries of coastal sediments incubated at 17 °C and amended with carbon in the form of chlorella algae and nitrate (50 μmol L−1) (Brin et al., 2015). However, the rates of potential DNRA observed by Brin et al. (2015) are an order of magnitude higher than those we observed in this study. Typically, the addition of either acetate or NOC did not significantly stimulate DNF and DNRA in the low nitrate treatments. However, rates of both DNF and DNRA increased with the addition of OC in the high nitrate treatments. Previous studies have

M.M. Rahman et al. / Science of the Total Environment 666 (2019) 608–617

2+

Huntingdale (+Fe )

Huntingdale (+Fe+Ac) 400

(b) 15NH + 4 Fe2+

250 200

400

150 100

200

600

300

400

200

200

100

15 NH + (µmo l/L ) 4

Fe 2+ 15NH + 4

600 F e 2 + (µmo l/L )

800

300

(a)

15 NH + (µmo l/L ) 4 F e 2 + (µmo l/L )

800

613

50 0

0

0 10

20

30

40

50

0 0

60

10

600 40 400

60

15NH + 4 Fe2+

80

800 60 600 40 400

20

200

20

200

0

0 10

20

30

40

50

0

60

0 0

10

Time (h) Koolamara (+Fe)

1200

30

40

50

60

Time (h)

Koolamara (+Fe+Ac) 1400

(e) 15NH + 4 Fe2+

80

800

60

600 40 400

50

(f)

15NH + 4 Fe2+

1200

15 NH + (µmo l/L ) 4 F e 2 + (µmo l/L )

1000

20

40

1000 30

800 600

20

400

20

10

200

200

0

0

0 10

20

30

40

50

0 0

60

10

40

50

60

Cascades (+Fe+Ac) 8

(g)

15NH + 4 Fe2+

6 800 4

600 400

2 200 0

0 10

20

30 Time (h)

40

50

60

30

(h)

1200

15NH + 4 Fe2+

25

1000 15 NH + (µmo l/L ) 4 F e 2 + (µmo l/L )

1000

0

30 Time (h)

Time (h) Cascades (+Fe) 1200

20

20 800 15

600 400

10

200

5

0

15 NH + (µmo l/L ) 4

0

F e 2 + (µmo l/L )

15 NH + (µmo l/L ) 4

0

F e 2 + (µmo l/L )

50

100

(d)

1000

60

15 NH + (µmo l/L ) 4 F e 2 + (µmo l/L )

F e 2 + (µmo l/L )

1200

15NH + 4 Fe2+

800

40

Namatjira (+Fe+Ac)

80

(c)

30 Time (h)

Time (h) Namatjira (+Fe)

1000

20

15 NH + (µmo l/L ) 4

0

0 0

10

20

30

40

50

60

Time (h)

Fig. 2. Decrease in added Fe2+ during pre-incubation before adding tracer (0 to ~25–30 h) and reduction of added Fe2+ with concomitant accumulation of 15NH+ 4 after tracer addition (~25–30 h to 60 h) from Huntingdale Road (a, b), Namatjira Reserve (c, d), Koolamara Blvd (e, f) and Cascades on Clyde (g, h) wetlands.

614

M.M. Rahman et al. / Science of the Total Environment 666 (2019) 608–617

DNF (µmo l L s lu r r y -1 h -1)

90

a

Control Fe Fe+Ac Ac

75 60 45

* * 30

*

15 * 0 Huntingdale

Namatjira

Koolamara

Cascades

Sites

DNRA (µmo l L s lu r r y -1 h -1)

90

*

b

Control Fe Fe+Ac Ac

*

75 60 45 30

*

* * *

15

* *

0 Huntingdale

Namatjira

Koolamara

*

Cascades

(Hardison et al., 2015; Kraft et al., 2014; van den Berg et al., 2016; Yoon et al., 2015). The addition of different types of OC did not make significant impact on either DNF or DNRA in the low nitrate treatments. However, the effect of OC addition in the high nitrate treatments was significant (p b .05) on both DNF and DNRA (Fig. 1). In 3 out of 4 wetlands, the significant (p b .05) decrease in the DNF:DNRA ratio at low OC/NO3− ratios suggests that addition of OC stimulated DNRA more than DNF (Fig. 1e and f). Moreover, in 3 out of 4 wetlands, the lowest DNF:DNRA ratios from the NOC treatments indicate that NOC stimulated DNRA more than acetate. Based on these findings we infer that the addition of NOC seems to stimulate fermentation and therefore DNRA because nitrate ammonifying bacteria are typically fermenters, which cannot further utilise acetate (Błaszczyk et al., 1980; Burgin and Hamilton, 2007). Błaszczyk et al. (1980) observed that in treatments with acetate and glucose, denitrifying bacteria accounted for 95 and 5% of the microbial population, respectively. It was demonstrated that fermentative bacteria competed for C in the glucose-amended treatment (Błaszczyk et al., 1980). In addition to types of OC, chemical composition of OC also plays an important role in partitioning between DNRA and DNF. Differences in the chemical composition between the added and native carbon were speculated as a factor in the trade-off between nitrate reduction to NH4+ or N2 (Akunna et al., 1993; Bonin et al., 1999; Gardner and McCarthy, 2009). For example, higher than expected DNRA rates were observed from shallow tropical sediments which was attributed to the composition of the OC (Gardner and McCarthy, 2009). Therefore, we suggest that higher potential rates of DNRA from the NOC treatment in this study are due to the similar chemical composition of the added NOC to the pre-existing sediment OC in slurries.

Sites

4.2. Iron-induced nitrate reduction

12

c

Control Fe Fe+Ac Ac

DNF : DNRA

10 8 6

* 4 * * *

2

* 0

*

* * * Huntingdale

Namatjira

Koolamara

Cascades

Sites Fig. 3. Rates of denitrification (a), DNRA (b) and their relative proportion (c) from slurries amended with acetate and Fe2+. Asterisks (*) above bars represent significant (p b .05, ANOVA) differences between control and acetate- and Fe2+-amended slurries. Error bars are ±1SD (n = 3).

shown that the ratio of electron donor (organic carbon) to electron acceptor (nitrate) is an important factor that often determines the partitioning between DNF and DNRA (Burgin and Hamilton, 2007; van den Berg et al., 2016; Yoon et al., 2015). van den Berg et al. (2016)) observed that DNRA bacteria dominated when nitrate was limited (OC/ NO3− = 1.87) while denitrifiers dominated under OC (acetate) limiting conditions (OC/NO3− = 0.66) in a chemostat culture. The added OC: NO3− ratios in our study were 0.2 and 4 from the high and low nitrate treatments, respectively. Although rates of both DNF and DNRA increased at the low OC/NO3− ratio, denitrifiers dominated at both high and low OC/NO3− ratios. Furthermore, an increase of 1.3 to 5.5 times in the DNF:DNRA ratio at low compared to high OC/NO3− ratios suggests that increased availability of nitrate stimulated DNF more than DNRA. This result supports our hypothesis of predominance of DNF under increased supply of nitrate and agrees with most previous studies

Our results showed that Fe2+ stimulated DNRA at the expense of DNF from all four wetlands. Rates of DNRA increased dramatically by 44–253% after the addition of Fe2+ into slurry samples. Similar results for different ecosystems were also observed from previous studies. For example, addition of Fe2+ to slurries of freshwater lake sediment (Lake Almind, Denmark) resulted in an increase of 76% in DNRA compared to parallel incubations without Fe2+ addition, and in some cases it also resulted in a concurrent decrease in DNF (Robertson and Thamdrup, 2017). A 50% increase in rates of DNRA was observed with increased availability of dissolved Fe2+ in the Yarra River estuary and an enhancement in DNRA at some study sites occurred at the expense of DNF (Robertson et al., 2016). Robertson and Thamdrup (2017) demonstrated that a small increase in the concentration of Fe2+ can cause a switch in the partitioning between DNF and DNRA in freshwater lake sediments. In our study, similar to Lake Almind, DNF was typically the dominant nitrate reduction pathway while a small portion of nitrate was reduced through DNRA when only nitrate was added to slurries. However, when Fe2+ was added into slurries together with nitrate, in 50% of cases, nitrate was predominantly reduced through DNRA. It was observed that during nitrate-dependent ferrous Fe oxidation, nitrate was reduced to NH4+ through DNRA by Candidatus Brocadia Sinica bacteria, which carries a gene essential for DNRA (nrfA-encoding ammonia-forming nitrite reductase) but does not contain a gene for DNF (nosZ, which encodes nitrous oxide reductase) (Oshiki et al., 2013). Furthermore, Weber et al. (2006b) and Coby et al. (2011) observed that due to the addition of nitrate, an immediate oxidation of Fe2+ occurred coupled to reduction of NO3− to NH4+ through DNRA in freshwater wetland and river sediments containing Geobacter sp., which are capable of oxidizing Fe2+ with concurrent reduction of NO3− to NH4+ (Lovley et al., 1993). Rates of DNRA from the acetate treatments are quite comparable to those from the Fe2+ + acetate treatments in most cases, which suggest that heterotrophic DNRA also concurrently occurred with the Fe2+ driven autotrophic DNRA. An increase of 1.5–4.5 times in the Fe2+

M.M. Rahman et al. / Science of the Total Environment 666 (2019) 608–617 8

6

4

2+

consumption rates

Fe Fe+Ac

DNRA:Fe

consumption rates after the addition of acetate indicates that Fe2+ driven DNRA progressed mixotrophically. Some studies have identified autotrophic nitrate-reducing iron-oxidizing strains (Li et al., 2014; Vorholt et al., 1997; Weber et al., 2006a) whereas most other strains carrying out DNRA are known to have mixotrophic (mixed lithotrophic and organotrophic) metabolism where they oxidize an organic substrate such as acetate during their growth (Benz et al., 1998; Chakraborty and Picardal, 2013; Straub et al., 1996). Although previous studies have shown sulfide can also act as an electron donor in addition to Fe2+ (Brettar and Rheinheimer, 1991; Brunet and Garcia-Gil, 1996; Burgin and Hamilton, 2008; Dong et al., 2011), we did not consider this due to the low sulfide concentrations (0.1–12 μmol L−1) in sediment porewater (Rahman et al., 2019). Some previous studies have shown that nitrate reduction to ammonium can occur via abiotic pathways (Hansen et al., 1996b; Klueglein and Kappler, 2013; Tai and Dempsey, 2009) and we cannot rule out that abiotic processes contribute to the rates observed here. However, we believe that nitratedependent Fe2+ oxidation in this study is a microbially mediated process since most previous studies have suggested abiotic Fe oxidation during nitrate reduction was negligible, including in the slurry approach used here (Klüber and Conrad, 1998; Robertson et al., 2016; Robertson and Thamdrup, 2017; Weber et al., 2001; Weber et al., 2006b). We also note that this study expands on the limited number of studies that use environmentally relevant concentrations of NO3− and Fe2+ (Robertson et al., 2016, Robertson and Thamdrup, 2017, Kessler et al., 2018), and confirms the importance of Fe2+ in the partitioning of DNF and DNRA at environmentally relevant concentrations. Nitrate-dependent Fe2+ oxidation may also stimulate DNF. The majority of bacterial isolates able to use Fe2+ as an electron donor to reduce nitrate were found to produce N2 as the end product (Chakraborty and Picardal, 2013; Muehe et al., 2009; Straub et al., 1996). However, addition of Fe2+ to slurries in our study resulted in a 18–39% decrease in DNF rates. Although the precise mechanism for this decrease in DNF is not clear, we suggest that DNRA bacteria stimulated by Fe2+ outcompeted DNF bacteria. A similar result was also observed from the sediments of the Yarra River estuary (Robertson et al., 2016) and Lake Almind (Robertson and Thamdrup, 2017). Of all wetlands, the DNRA rates were highest from the Huntingdale Road wetland, which had the highest dissolved Fe2+ concentrations when nitrate was added (Fig. 2b). In contrast, the lowest DNRA rates were observed from the Cascades on Clyde wetland, which had the lowest dissolved Fe2+ concentrations at the time of nitrate addition (Fig. 2h). These results further demonstrate that increased availability of Fe2+ stimulated the rates of DNRA. The stoichiometries of rates of DNRA to rates of Fe2+ consumption were somewhat less (typically around 2–5 instead of the expected 8) compared to those expected for Fe2+-stimulated nitrate reduction to ammonium (Eq. (2), Fig. 4). However, the stoichiometries were close to those expected for Fe2+-stimulated nitrite reduction to ammonium (Eq. (4), Fig. 4). The ratio of DNRA to Fe consumption was nearly twice as high as expected values for freshwater wetlands and rivers (Coby et al., 2011; Weber et al., 2006b) and the Yarra River estuary (Roberts et al., 2014); the ratio for Lake Almind corresponded to the expected value of nitrate reduction to ammonium production (Robertson and Thamdrup, 2017). There may be several reasons for the imbalance between the stoichiometries of rates of DNRA to rates of Fe2+ consumption: 1. NO3− could initially be organotrophically reduced to NO2−, in which OC acts as the electron donor; subsequently, lithotrophic, nitrite-dependent Fe2+ oxidation coupled to DNRA may occur (Coby et al., 2011; Weber et al., 2006b). In support of this hypothesis, Robertson et al. (2016) observed an Fe2+ oxidation to DNRA ratio of 5.3, which was very close to the predicted stoichiometry of 6 from their nitrite experiments on the Yarra River estuary; 2. Alternatively, since most of the added Fe2+ was bound to sediment particles, there was some desorption of particlebound Fe(II) as evident from both the Fe2+ and Fe2+ + acetate

615

2

0 Huntingdale

Namatjira

Koolamara

Cascades

Sites

Fig. 4. Stoichiometry between the rates of Fe2+ consumption and 15NH+ 4 production from slurries amended with Fe2+ and Fe2+ + acetate.

treatments of the Huntingdale Road wetland (Fig. 2a and b), which eventually resulted in a lower ratio of Fe2+ oxidized to DNRA than predicted. 5. Implication for management This study showed that addition of the native OC into the sediments of these wetlands promoted both removal (DNF) and recycling (DNRA) of NO3− when the concentration of NO3− was high. However, the addition of NOC led to an increase in the recycling of nitrogen relative to removal. Therefore, the management strategy of these constructed wetlands should carefully consider the types and densities of the macrophytes, for example, Typha spp. and Phragmites australis introduced in them. Moreover, regular harvesting of the vegetation to reduce the decomposition of plant litter may improve the N removal capacity of constructed wetlands. Furthermore, the addition of Fe2+ in slurries increased the recycling rates of NO3− from 40 to 250%, which supports the finding of Fe2+ linked to DNRA. The presence of high Fe2+ in sediment porewater may also reduce the N removal capacity of constructed wetlands by stimulating the recycling of N. Most of the reactive iron in wetlands probably originates from the surface of particulate inorganic matter such as silts and clays. Fortuitously, a common design feature of wetlands is a sediment trap at the entrance to the wetland, which would lead to the removal of most of the iron entering the system (Kadlec and Wallace, 2009). Other factors that control the amount of iron in the wetland sediments include inputs of iron rich groundwater, as well as the catchment geology. Further research is required to elucidate if there is any relationship between nitrogen removal efficiency in wetlands and their sediment organic carbon and Fe2+ content. 6. Conclusions We investigated the role of organic carbon and nitrate at two different OC/NO3− ratios and of Fe2+ on the partitioning of nitrate reduction between DNF and DNRA in sediments of four constructed stormwater urban wetlands. Our results suggest that DNF was the dominant nitrate reduction process regardless of the OC/NO3− ratios. However, rates of both DNF and DNRA increased at elevated nitrate concentrations, which further suggests that nitrate reduction in these sediments was limited by nitrate supply. In most cases, increased nitrate concentrations led to a stimulation of DNF over DNRA, consistent with previous findings in these wetlands (Rahman et al., 2019). Higher rates of both DNF and DNRA were observed when NOC was added compared to

616

M.M. Rahman et al. / Science of the Total Environment 666 (2019) 608–617

acetate suggesting that NOC was preferable to both denitrifying and nitrate ammonifying bacteria. Our results from slurry incubations with Fe2+ indicate that Fe2+ had a negative effect on DNF and a positive influence on DNRA. An enhancement in DNF rates with acetate addition suggests that DNF in this study was predominantly heterotrophic. In contrast, higher rates of DNRA from the Fe2+ and acetate additions (compared to the control) suggest that unlike DNF, DNRA occurred both autotrophically and heterotrophically. This strong link of Fe2+ with DNRA is also supported by a previous study on these wetlands (Rahman et al., 2019), which showed that DNRA controlled the porewater Fe2+ pool size. Therefore, our results support the hypothesis that increased availability of Fe2+ leads to recycling of N through increased DNRA in sediments and thus, Fe2+ plays a significant role in altering the relative importance of N cycling processes in constructed stormwater urban wetlands. 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 discussions on the Koolamara Blvd Wetland. Jesse Pottage, Dale Christensen, Douglas Russell, Michael Bourke, David Brehm, Wei Wen Wong and Bipasa Akter are all thanked for their assistance with fieldwork. Md. Moklesur Rahman is grateful to the Cooperative Research Centre for Water Sensitive Cities for funding support in carrying out this project and Melbourne Water for providing background information on the wetlands. This work was also supported by the Australian Research Council grant DP150101281 to PLMC. Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi. org/10.1016/j.scitotenv.2019.02.225. References Achtnich, C., Bak, F., Conrad, R., 1995. Competition for electron donors among nitrate reducers, ferric iron reducers, sulfate reducers, and methanogens in anoxic paddy soil. Biol. Fertil. Soils 19, 65–72. Akunna, J.C., Bizeau, C., Moletta, R., 1993. Nitrate and nitrite reductions with anaerobic sludge using various carbon sources: glucose, glycerol, acetic acid, lactic acid and methanol. Water Res. 27, 1303–1312. Benz, M., Brune, A., Schink, B., 1998. Anaerobic and aerobic oxidation of ferrous iron at neutral pH by chemoheterotrophic nitrate-reducing bacteria. Arch. Microbiol. 169, 159–165. Błaszczyk, M., Mycielski, R., Jaworowska-Deptuch, H., Brzostek, K., 1980. Effect of various sources of organic carbon and high nitrite and nitrate concentrations on the selection of denitrifying bacteria. I. Stationary cultures. Acta Microbiol. Pol. 29, 397. Bonaglia, S., Klawonn, I., De Brabandere, L., Deutsch, B., Thamdrup, B., Brüchert, V., 2016. Denitrification and DNRA at the Baltic Sea oxic–anoxic interface: substrate spectrum and kinetics. Limnol. Oceanogr. 61, 1900–1915. Bonin, P., 1996. Anaerobic nitrate reduction to ammonium in two strains isolated from coastal marine sediment: a dissimilatory pathway. FEMS Microbiol. Ecol. 19, 27–38. Bonin, P., Omnes, P., Chalamet, A., 1999. The influence of nitrate and carbon inputs on the end products of bacterial nitrate dissimilation in marine sediment. Toxicol. Environ. Chem. 73, 67–79. Brettar, I., Rheinheimer, G., 1991. Denitrification in the Central Baltic: evidence for H2Soxidation as motor of denitrification at the oxic-anoxic interface. Mar. Ecol. Prog. Ser. 77, 157–169. Brin, L.D., Giblin, A.E., Rich, J.J., 2015. Effects of experimental warming and carbon addition on nitrate reduction and respiration in coastal sediments. Biogeochemistry 125, 81–95. Brunet, R.C., Garcia-Gil, L.J., 1996. Sulfide-induced dissimilatory nitrate reduction to ammonia in anaerobic freshwater sediments. FEMS Microbiol. Ecol. 21, 131–138. Burgin, A.J., Hamilton, S.K., 2007. Have we overemphasized the role of denitrification in aquatic ecosystems? A review of nitrate removal pathways. Front. Ecol. Environ. 5, 89–96. Burgin, A.J., Hamilton, S.K., 2008. NO3−-driven SO42− production in freshwater ecosystems: implications for N and S cycling. Ecosystems 11, 908–922. Chakraborty, A., Picardal, F., 2013. Neutrophilic, nitrate-dependent, Fe (II) oxidation by a Dechloromonas species. World J. Microbiol. Biotechnol. 29, 617.

Coby, A.J., Picardal, F., Shelobolina, E., Xu, H., Roden, E.E., 2011. Repeated anaerobic microbial redox cycling of iron. Appl. Environ. Microbiol. 77, 6036–6042. Dalsgaard, T., Nielsen, L.P., Brotas, V., Viaroli, P., Underwood, G., Nedwell, D., et al., 2000. Protocol Handbook for NICE-Nitrogen Cycling in Estuaries: A Project Under the EU Research Programme: Marine Science and Technology (MAST III). National Environmental Research Institute, Denmark. DeLaune, R.D., Pezeshki, S.R., Jugsujinda, A., 2005. Impact of Mississippi River freshwater reintroduction on Spartina patens marshes: responses to nutrient input and lowering of salinity. Wetlands 25, 155–161. Devlin, J.F., Eedy, R., Butler, B.J., 2000. The effects of electron donor and granular iron on nitrate transformation rates in sediments from a municipal water supply aquifer. J. Contam. Hydrol. 46, 81–97. Dodla, S.K., Wang, J.J., DeLaune, R.D., Cook, R.L., 2008. Denitrification potential and its relation to organic carbon quality in three coastal wetland soils. Sci. Total Environ. 407, 471–480. Dong, L.F., Sobey, M.N., Smith, C.J., Rusmana, I., Phillips, W., Stott, A., et al., 2011. Dissimilatory reduction of nitrate to ammonium, not denitrification or anammox, dominates benthic nitrate reduction in tropical estuaries. Limnol. Oceanogr. 56, 279–291. Emerson, D., Weiss, J.V., Megonigal, J.P., 1999. Iron-oxidizing bacteria are associated with ferric hydroxide precipitates (Fe-plaque) on the roots of wetland plants. Appl. Environ. Microbiol. 65, 2758–2761. Finneran, K.T., Housewright, M.E., Lovley, D.R., 2002. Multiple influences of nitrate on uranium solubility during bioremediation of uranium-contaminated subsurface sediments. Environ. Microbiol. 4, 510–516. Fossing, H., Gallardo, V.A., Jorgensen, B.B., Huttel, M., Nielsen, L.P., Schulz, H., et al., 1995. Concentration and transport of nitrate by the mat-forming sulphur bacterium Thioploca. Nature 374, 713–715. Gardner, W., McCarthy, M., 2009. Nitrogen dynamics at the sediment–water interface in shallow, sub-tropical Florida Bay: why denitrification efficiency may decrease with increased eutrophication. Biogeochemistry 95, 185–198. Greenan, C.M., Moorman, T.B., Kaspar, T.C., Parkin, T.B., Jaynes, D.B., 2006. Comparing carbon substrates for denitrification of subsurface drainage water. J. Environ. Qual. 35, 824–829. Hansen, H.C.B., Koch, C.B., Nancke-Krogh, H., Borggaard, O.K., Sørensen, J., 1996a. Abiotic nitrate reduction to ammonium: key role of green rust. Environ. Sci. Technol. 30, 2053–2056. Hansen, H.C.B., Koch, C.B., Nancke-Krogh, H., Borggaard, O.K., Sørensen, J., 1996b. Abiotic nitrate reduction to ammonium: key role of green rust. Environ. Sci. Technol. 30, 2053–2056. Hardison, A.K., Algar, C.K., Giblin, A.E., Rich, J.J., 2015. Influence of organic carbon and nitrate loading on partitioning between dissimilatory nitrate reduction to ammonium (DNRA) and N2 production. Geochim. Cosmochim. Acta 164, 146–160. Her, J.-J., Huang, J.-S., 1995. Influences of carbon source and C/N ratio on nitrate/nitrite denitrification and carbon breakthrough. Bioresour. Technol. 54, 45–51. Hou, L., Liu, M., Carini, S.A., Gardner, W.S., 2012. Transformation and fate of nitrate near the sediment–water interface of Copano Bay. Cont. Shelf Res. 35, 86–94. Kadlec, R.H., Wallace, S.D., 2009. Treatment Wetlands. Kappler, A., Straub, K.L., 2005. Geomicrobiological cycling of iron. Rev. Mineral. Geochem. 59, 85–108. Kelly-Gerreyn, B.A., Trimmer, M., Hydes, D.J., 2001. A diagenetic model discriminating denitrification and dissimilatory nitrate reduction to ammonium in a temperate estuarine sediment. Mar. Ecol. Prog. Ser. 220, 33–46. Kelso, B., Smith, R.V., Laughlin, R.J., Lennox, S.D., 1997. Dissimilatory nitrate reduction in anaerobic sediments leading to river nitrite accumulation. Appl. Environ. Microbiol. 63, 4679–4685. Kessler, A.J., Roberts, K.L., Bissett, A., Cook, P.L.M., 2018. Biogeochemical controls on the relative importance of denitrification and dissimilatory nitrate reduction to ammonium in estuaries. Glob. Biogeochem. Cycles 32, 1045–1057. https://doi.org/ 10.1029/2018GB005908. Klüber, H.D., Conrad, R., 1998. Effects of nitrate, nitrite, NO and N2O on methanogenesis and other redox processes in anoxic rice field soil. FEMS Microbiol. Ecol. 25, 301–318. Klueglein, N., Kappler, A., 2013. Abiotic oxidation of Fe(II) by reactive nitrogen species in cultures of the nitrate-reducing Fe(II) oxidizer Acidovorax sp. BoFeN1 - questioning the existence of enzymatic Fe(II) oxidation. Geobiology 11, 180–190. Kraft, B., Tegetmeyer, H.E., Sharma, R., Klotz, M.G., Ferdelman, T.G., Hettich, R.L., et al., 2014. The environmental controls that govern the end product of bacterial nitrate respiration. Science 345, 676–679. Krishna Mohan, T.V., Nancharaiah, Y.V., Venugopalan, V.P., Satya Sai, P.M., 2016. Effect of C/N ratio on denitrification of high-strength nitrate wastewater in anoxic granular sludge sequencing batch reactors. Ecol. Eng. 91, 441–448. Laufer, K., Røy, H., Jørgensen, B.B., Kappler, A., 2016. Evidence for the existence of autotrophic nitrate-reducing Fe (II)-oxidizing bacteria in marine coastal sediment. Appl. Environ. Microbiol. 82, 6120–6131. Li, B., Tian, C., Zhang, D., Pan, X., 2014. Anaerobic nitrate-dependent iron (II) oxidation by a novel autotrophic bacterium, Citrobacter freundii strain PXL1. Geomicrobiol J. 31, 138–144. Liu, X., Han, J.-G., Ma, Z.-W., Wang, Q., Li, L.-H., 2016. Effect of carbon source on dissimilatory nitrate reduction to ammonium in Costal Wetland sediments. J. Soil Sci. Plant Nutr. 16, 337–349. Lovley, D.R., Giovannoni, S., White, D., Champine, J., Phillips, E., Gorby, Y., et al., 1993. Geobacter metallireducens gen. nov. sp. nov., a microorganism capable of coupling the complete oxidation of organic compounds to the reduction of iron and other metals. Arch. Microbiol. 159, 336–344. Melton, E.D., Schmidt, C., Kappler, A., 2012. Microbial iron (II) oxidation in littoral freshwater lake sediment: the potential for competition between phototrophic vs. nitrate-reducing iron (II)-oxidizers. Front. Microbiol. 3.

M.M. Rahman et al. / Science of the Total Environment 666 (2019) 608–617 Melton, E.D., Stief, P., Behrens, S., Kappler, A., Schmidt, C., 2014. High spatial resolution of distribution and interconnections between Fe-and N-redox processes in profundal lake sediments. Environ. Microbiol. 16, 3287–3303. Muehe, E.M., Gerhardt, S., Schink, B., Kappler, A., 2009. Ecophysiology and the energetic benefit of mixotrophic Fe(II) oxidation by various strains of nitratereducing bacteria. FEMS Microbiol. Ecol. 70, 335–343. Nielsen, L.P., 1992. Denitrification in sediment determined from nitrogen isotope pairing. FEMS Microbiol. Lett. 86, 357–362. Nijburg, J.W., Coolen, M., Gerards, S., Gunnewiek, P., Laanbroek, H.J., 1997. Effects of nitrate availability and the presence of Glyceria maxima on the composition and activity of the dissimilatory nitrate-reducing bacterial community. Appl. Environ. Microbiol. 63, 931–937. Oshiki, M., Ishii, S., Yoshida, K., Fujii, N., Ishiguro, M., Satoh, H., et al., 2013. Nitratedependent ferrous iron oxidation by anaerobic ammonium oxidation (Anammox) bacteria. Appl. Environ. Microbiol. 79, 4087–4093. Paul, J.W., Beauchamp, E.G., Trevors, J.T., 1989. Acetate, propionate, butyrate, glucose, and sucrose as carbon sources for denitrifying bacteria in soil. Can. J. Microbiol. 35, 754–759. Rahman, M.M., Roberts, K.L., Warry, F., Grace, M.R., Cook, P.L.M., 2019. Factors controlling dissimilatory nitrate reduction processes in constructed stormwater urban wetlands. Biogeochemistry 142, 375–393. Reyes-Avila, J., Razo-Flores, Ea, Gomez, J., 2004. Simultaneous biological removal of nitrogen, carbon and sulfur by denitrification. Water Res. 38, 3313–3321. Rich, J.J., Dale, O.R., Song, B., Ward, B.B., 2008. Anaerobic ammonium oxidation (anammox) in Chesapeake Bay sediments. Microb. Ecol. 55, 311–320. Risgaard-Petersen, N., Revsbech, N.P., Rysgaard, S., 1995. Combined microdiffusionhypobromite oxidation method for determining Nitrogen-15 isotope in ammonium. Soil Sci. Soc. Am. J. 59, 1077–1080. Roberts, K.L., Kessler, A.J., Grace, M.R., Cook, P.L.M., 2014. Increased rates of dissimilatory nitrate reduction to ammonium (DNRA) under oxic conditions in a periodically hypoxic estuary. Geochim. Cosmochim. Acta 133, 313–324. Robertson, E.K., Thamdrup, B., 2017. The fate of nitrogen is linked to iron(II) availability in a freshwater lake sediment. Geochim. Cosmochim. Acta 205, 84–99. Robertson, E.K., Roberts, K.L., Burdorf, L.D., Cook, P., Thamdrup, B., 2016. Dissimilatory nitrate reduction to ammonium coupled to Fe (II) oxidation in sediments of a periodically hypoxic estuary. Limnol. Oceanogr. 61, 365–381. Schmidt, C.S., Richardson, D.J., Baggs, E.M., 2011. Constraining the conditions conducive to dissimilatory nitrate reduction to ammonium in temperate arable soils. Soil Biol. Biochem. 43, 1607–1611. Sgouridis, F., Heppell, C.M., Wharton, G., Lansdown, K., Trimmer, M., 2011. Denitrification and dissimilatory nitrate reduction to ammonium (DNRA) in a temperate reconnected floodplain. Water Res. 45, 4909–4922. Stevens, R.J., Laughlin, R.J., Malone, J.P., 1998. Soil pH affects the processes reducing nitrate to nitrous oxide and di-nitrogen. Soil Biol. Biochem. 30, 1119–1126. Straub, K.L., Buchholz-Cleven, B.E.E., 1998. Enumeration and detection of anaerobic ferrous iron-oxidizing, nitrate-reducing bacteria from diverse European sediments. Appl. Environ. Microbiol. 64, 4846–4856.

617

Straub, K.L., Benz, M., Schink, B., Widdel, F., 1996. Anaerobic, nitrate-dependent microbial oxidation of ferrous iron. Appl. Environ. Microbiol. 62, 1458–1460. Straub, K.L., Schönhuber, W.A., Buchholz-Cleven, B.E.E., Schink, B., 2004. Diversity of ferrous iron-oxidizing, nitrate-reducing bacteria and their involvement in oxygenindependent Iron cycling. Geomicrobiol J. 21, 371–378. Strohm, T.O., Griffin, B., Zumft, W.G., Schink, B., 2007. Growth yields in bacterial denitrification and nitrate ammonification. Appl. Environ. Microbiol. 73, 1420–1424. Tai, Y.-L., Dempsey, B.A., 2009. Nitrite reduction with hydrous ferric oxide and Fe(II): stoichiometry, rate, and mechanism. Water Res. 43, 546–552. Tam, N.F.Y., Wong, Y.S., Leung, G., 1992. Effect of exogenous carbon sources on removal of inorganic nutrient by the nitrification-denitrification process. Water Res. 26, 1229–1236. Tiedje, J.M., 1988. Ecology of denitrification and dissimilatory nitrate reduction to ammonium. Biol. Anaerobic Microorg. 717, 179–244. Tiedje, J.M., Sexstone, A.J., Myrold, D.D., Robinson, J.A., 1983. Denitrification: ecological niches, competition and survival. Antonie Van Leeuwenhoek 48, 569–583. Tomaszek, J.A., Rokosz, G.R., 2007. Rates of dissimilatory nitrate reduction to ammonium in two polish reservoirs: impacts of temperature, organic matter content, and nitrate concentration. Environ. Technol. 28, 771–778. van den Berg, E.M., Boleij, M., Kuenen, J.G., Kleerebezem, R., van Loosdrecht, M.C.M., 2016. DNRA and denitrification coexist over a broad range of acetate/N-NO(3)(−) ratios, in a chemostat enrichment culture. Front. Microbiol. 7, 1842. Vorholt, J.A., Hafenbradl, D., Stetter, K.O., Thauer, R.K., 1997. Pathways of autotrophic CO2 fixation and of dissimilatory nitrate reduction to N2O in Ferroglobus placidus. Arch. Microbiol. 167, 19–23. Weber, K.A., Picardal, F.W., Roden, E.E., 2001. Microbially catalyzed nitrate-dependent oxidation of biogenic solid-phase Fe(II) compounds. Environ. Sci. Technol. 35, 1644–1650. Weber, K.A., Pollock, J., Cole, K.A., O'Connor, S.M., Achenbach, L.A., Coates, J.D., 2006a. Anaerobic nitrate-dependent iron (II) bio-oxidation by a novel lithoautotrophic betaproteobacterium, strain 2002. Appl. Environ. Microbiol. 72, 686–694. Weber, K.A., Urrutia, M.M., Churchill, P.F., Kukkadapu, R.K., Roden, E.E., 2006b. Anaerobic redox cycling of iron by freshwater sediment microorganisms. Environ. Microbiol. 8, 100–113. White, J.R., Reddy, K., 2003. Nitrification and denitrification rates of Everglades wetland soils along a phosphorus-impacted gradient. J. Environ. Qual. 32, 2436–2443. Yin, S.X., Chen, D., Chen, L.M., Edis, R., 2002. Dissimilatory nitrate reduction to ammonium and responsible microorganisms in two Chinese and Australian paddy soils. Soil Biol. Biochem. 34, 1131–1137. Yoon, S., Cruz-Garcia, C., Sanford, R., Ritalahti, K.M., Loffler, F.E., 2015. Denitrification versus respiratory ammonification: environmental controls of two competing dissimilatory NO3-/NO2- reduction pathways in Shewanella loihica strain PV-4. ISME J. 9, 1093–1104.