Denitrification and benthic metabolism in lowland pit lakes: The role of trophic conditions

Science of the Total Environment xxx (xxxx) xxx

Contents lists available at ScienceDirect

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

Denitrification and benthic metabolism in lowland pit lakes: The role of trophic conditions Daniele Nizzoli a,⇑, David T. Welsh b, Pierluigi Viaroli a a b

Department of Chemistry, Life Sciences and Environmental Sustainability, University of Parma, Parco Area delle Scienze 11/A, 43124 Parma, Italy School of Environment and Environmental Futures Research Institute, Griffith University, Gold Coast Campus, PMB 50 GC Mail Centre, Bundall 9726, Queensland, 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


N-cycling rates were measured in

hypolimnetic and littoral sediments of 5 pit lakes.
Sediments were net N-sinks in oligomesotrophic, but N-sources in eutrophic lakes.
Denitrification accounted for 60 & 96% of N-removal in littoral and hypolimnetic sediments.
Pelagic assimilation became the dominant N-sink with increasing eutrophication status.
If properly managed, pit lakes can replace lost wetlands as nutrient filters.

a r t i c l e

i n f o

Article history: Received 2 July 2019 Received in revised form 2 October 2019 Accepted 2 October 2019 Available online xxxx Editor: Sergi Sabater Keywords: Benthic nitrogen fluxes Denitrification Nitrogen removal Sand pit lakes Restoration Trophic state

a b s t r a c t Over recent decades, a great number of pit lakes have been formed, as a result of sand and gravel quarrying in river floodplains that are often also heavily exploited for agriculture. These lakes can act as nutrient filters and regulate the nitrogen pollution resulting from agricultural fertiliser use. In this paper we report the main outcomes of a study of the major nitrogen pathways in five pit lakes of differing trophic status, located along a lowland stretch of the Po river (Northern Italy). Benthic nitrogen fluxes and denitrification rates were determined in the hypolimnion and denitrification and reactive nitrogen assimilation by microphytobenthos in the littoral zone. We tested the hypothesis that lake depth and trophic status can impair denitrification and/or reactive nitrogen assimilation, compromising the function of the lakes as nutrient filters. In the studied lakes, denitrification and reactive nitrogen assimilation by primary producer communities accounted for substantial nitrogen removal rates, which were among the highest reported in the literature. Benthic nitrogen fluxes and denitrification varied between and within lakes, with depth. The littoral zone and surface waters also supported primary production, favouring nitrogen assimilation and temporal retention in the primary producer biomass. In all lakes, denitrification rates decreased from littoral to hypolimnetic sites. Denitrification rates and net nitrogen assimilation also diminished from oligotrophic to eutrophic conditions. To some extent, in eutrophic lakes there was a transfer of primary production from the benthos to the water column and the benthic system became heterotrophic, reducing the capacity for net nitrogen removal.

⇑ Corresponding author at: Department of Chemistry, Life Sciences and Environmental Sustainability, Parma University, Parco Area delle Scienze 11A, I-43124 Parma, Italy. E-mail address: [email protected] (D. Nizzoli). https://doi.org/10.1016/j.scitotenv.2019.134804 0048-9697/Ó 2019 Elsevier B.V. All rights reserved.

Please cite this article as: D. Nizzoli, D. T. Welsh and P. Viaroli, Denitrification and benthic metabolism in lowland pit lakes: The role of trophic conditions, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134804

2

D. Nizzoli et al. / Science of the Total Environment xxx (xxxx) xxx

Overall these results highlight that floodplain pit lakes can provide ecosystem services formerly supplied by natural wetlands. An important factor for management is the development of extensive littoral and shallow water zones, which are critical for maximising the nitrogen removal. Ó 2019 Elsevier B.V. All rights reserved.

1. Introduction Anthropogenic nitrogen (N) loads to watersheds increase reactive N (Nr) delivery to surface and ground waters (Galloway et al., 2003; Swaney et al., 2012) with severe negative effects, such as eutrophication of transitional and coastal marine waters (Howarth and Marino, 2006), alteration of aquatic food webs (Glibert, 2017) and human disease (Townsend et al., 2003; Schullehner et al., 2018). Additionally, the widespread destruction of pristine wetlands for urban and agricultural development, and the deterioration of natural aquatic ecosystems can greatly contribute to Nr pollution to water systems. Networks of streams, lakes, reservoirs, and wetlands are not simple conduits for waters, but play an important role in processing Nr loads from agricultural and urban areas (Nizzoli et al., 2018; Finlay et al., 2013; Seitzinger et al., 2006; Cheng and Basu, 2017). Exploitation of gravel and sand deposits along rivers and streams is a widespread practise to meet the needs for urban and infrastructure construction (Mollema and Antonellini, 2016). In Europe the demand for sand and gravel is estimated to be ~109 m3 per annum, with an associated turnover of 15 billion euros, and employment of ~200,000 workers (UEPG, 2017). In Italy, approximately 68  106 tons of sand and gravel were mined in 2014, of which ~60% was extracted from the Po river floodplain, corresponding to approx. 300 m3 per km2 y1 (National Institute of Statistic, 2017). Pit lakes are a result of this activity, especially when the quarry intercepts the water table or is connected to rivers and streams (Mollema and Antonellini, 2016). These aquatic ecosystems are becoming increasingly relevant for environmental restoration and management, due to their increasing number, and their shape and function, which are similar to riverine wetlands and small lakes with surface areas up to 100 ha and depths up to 20 m (Kattner et al., 2000; Weilhartner et al., 2012; Muellegger et al., 2013; Peckenham et al., 2009; Søndergaard et al., 2018). After cessation of extraction activities, pit lakes can evolve into valuable aquatic ecosystems that can be managed for reconstruction of aquatic habitats and ecological networks. In heavily impacted floodplains, they can take the place of lost natural components of the river margins such as, oxbow and riverine lakes, and ponds and temporary pools, providing key processes and functions, which regulate nutrient cycles (Søndergaard et al., 2018). In recent years, small, shallow lakes and secondary hydrographic networks, e.g. drainage ditches in croplands, have been increasingly recognized as valuable sinks for Nr (Seitzinger et al., 2006; Harrison et al., 2009; Castaldelli et al., 2015; Cheng and Basu, 2017). However, whether pit lakes can provide a similar function is still an open question, as they are so far understudied. Moreover, the magnitude of nitrogen removal, and the factors controlling Nr cycling pathways and the fate of Nr in these water bodies are poorly understood (Nizzoli et al., 2010; Blanchette and Lund, 2016). In addition, in pit lakes, Nr removal can be managed to some extent by designing and shaping the form of the lake. It is therefore important that the factors responsible for regulating Nr removal, retention and recycling are assessed and understood, in order to facilitate the design and management of these waterbodies to maximise ecological functions, such as the mitigation of Nr loads (Nizzoli et al., 2010, 2014).

The sediment-water interface constitutes an important compartment for Nr cycling. The array of processes such as sedimentation and accumulation, microbial ammonification, nitrification, denitrification, dissimilatory nitrate reduction to ammonium (DNRA), and the assimilation of Nr by primary producers in this zone are almost exclusively responsible for the benthic nitrogen cycle (Burgin and Hamilton, 2007; Nizzoli et al., 2014). The fate of Nr (i.e. whether Nr is permanently lost as N2, accumulated in the sediment, or recycled due to ammonification or DNRA) crucially depends upon the balance between these processes. Inlake conditions such as concentrations of dissolved oxygen (DO) and nitrate (NO 3 ), organic matter availability, primary production and bioturbation primarily control the benthic N cycling (Burgin and Hamilton, 2007; Finlay et al., 2013). Several combinations of these conditions arise from factors that change over space and time due to watershed management practises (e.g. land use and nutrient loads). For example, urbanization and agriculture increase allochthonous and/or autochthonous carbon inputs to the benthic system, as a result of increased soil erosion and runoff, and increase nutrient loading that stimulates within lake productivity and the potential for NO 3 removal (Bruesewitz et al., 2011). The efficiency of Nr removal is also related to trophic status, with Nr uptake being three times greater in eutrophic compared to oligotrophic lakes (Finlay et al., 2013). Benthic metabolism and Nr processing can also be highly variable within lakes (den Heyer and Kalff, 1998; Saunders and Kalff, 2001). For example, shallow littoral areas of lakes are considered as ‘‘hot spots” of denitrification, while benthic denitrification rates decrease with increasing water depth (Bruesewitz et al., 2012; Nizzoli et al., 2018). However, in shallow waters light availability can favour the development of benthic microalgal or macrophyte communities with contrasting effects on microbial Nr transformations (Dunn et al., 2012; Nizzoli et al., 2014; Risgaard-Petersen, 2003; Sundbäck et al., 2000). The biomass and activity of benthic microalgae is also affected by trophic state (Vadeboncoeur et al., 2003), with cascading effects on nitrogen cycling. In heavily exploited watersheds, high inputs of NO 3 from croplands and/or urban areas to pit lakes would be expected to stimulate NO 3 removal. However, such lakes would also be expected to undergo a rapid ageing process, shifting from oligotrophic to eutrophic conditions (Tavernini et al., 2009). We hypothesise that increased eutrophication status and/or the shift from net autotrophic to heterotrophic conditions would also shift the ratio between denitrification and Nr assimilation towards assimilation, which would impair the capacity of the lakes to mitigate Nr loads. To test this hypothesis, we performed a study in five pit lakes, selected as model ecosystems, to evaluate how benthic metabolism and related Nr processes varied under different trophic conditions and to explore which factors primarily regulated Nr removal and recycling. Specifically, to what extent denitrification varied among lakes, in relation to trophic status, and the degree of benthic autotrophy and heterotrophy, by evaluating 1) whether the benthic system was a net source or sink for Nr; 2) the extent of benthic denitrification and net Nr removal; 3) the degree that denitrification rates and benthic nitrogen fluxes varied within the lake and the factors influencing these variations, e.g. littoral versus hypolimnetic sediments; in each of the selected lakes.

Please cite this article as: D. Nizzoli, D. T. Welsh and P. Viaroli, Denitrification and benthic metabolism in lowland pit lakes: The role of trophic conditions, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134804

3

D. Nizzoli et al. / Science of the Total Environment xxx (xxxx) xxx

Fig. 1. Map of the Po river watershed showing areas of natural, agricultural and urban land use (upper panel) and the location of the five study lakes (lower panel).

2. Material and methods 2.1. Study site The study was carried out in five lakes formed by sand quarries close to the main course of the Po River (Northern Italy, Fig. 1) in an area of the floodplain that is heavily exploited for crop and livestock production (Viaroli et al., 2018). Approximately 80% of the area is cultivated for crops, while cattle (61 ind km2) and pigs (109 ind km2) are the main livestock species. Less than 10% of land surface is occupied by urban areas (Fig. 1). The selected lakes had maximum depths of 13 to 21 m and surface areas of 3.5 to 28 ha (Table 1). Three of the lakes, Ca’ Morta (CM), Ca’ Stanga (CS) and Lago Verde (LV) without in- or outlets received only direct precipitation, groundwater and overland runoff. Two lakes, Isola Giarola (IG) and Bosco della Lite (BL) are intermittently connected to the Po river during flood events. Persistent thermal stratification occurs between April and October in all 5

lakes, with water overturn taking place in November. The lakes differed in trophic status, with CM and CS being oligo-mesotrophic, LV meso-eutrophic and IG and BL hyper-eutrophic (Tavernini et al., 2009; Nizzoli et al., 2010). The littoral zone of the lakes was sparsely colonized by submerged vegetation dominated by Characeae (Chara cfr. fragilis e C. cfr. connivens) and species of the genus Potamogeton (Potamogeton pectinatus, P. lucens, P. natans e P. polygonifolius). This study was performed from May to July 2009. In each lake, samples were collected in two areas, the littoral zone (depth < 3 m; littoral sites) and a second site corresponding to the maximum depth point of each lake (hypolimnetic sites). 2.2. Water sampling and analyses At littoral and hypolimnetic sites, water temperature and DO were determined on site with a multiparameter probe (YSI 556). At the hypolimnetic sites, 2 L water samples were collected with

Table 1 Morphometric characteristics of the investigated pit lakes (CM = Ca0 Morta, CS = Ca0 Stanga, LV = Lago Verde, IG = Isola Giarola, BL = Bosco della Lite). Lake

Volume (m2)

Zmax (m)

Zm (m)

Shore (m)

Surface (m2)

Surface Littoral zone (m2)

Littoral zone/Lake surface (%)

CM CS LV IG BL

2,276,334 1,018,935 181,159 1,000,518 2,495,895

21 17 13 15 14

11 9 5 5 9

1875 1441 845 2329 2886

206,500 113,535 35,322 208,546 281,074

19,000 13,000 8000 50,000 32,000

9 11 23 24 11

Please cite this article as: D. Nizzoli, D. T. Welsh and P. Viaroli, Denitrification and benthic metabolism in lowland pit lakes: The role of trophic conditions, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134804

4

D. Nizzoli et al. / Science of the Total Environment xxx (xxxx) xxx

a Ruttner bottle at surface, 1, 2, 4, 6, 8, 9, 10 and 12 m depths, and ~50 cm above the sediment. At the littoral sites water was sampled only at ~50 cm above the sediment. Water samples were filtered (Whatman GF/F, nominal pore size 0.7 mm) and analysed within 24 h for ammonium (NH+4, Koroleff, 1970) and NO 3 (APHA, 1998). Total phosphorus was analysed on unfiltered water samples (Valderrama, 1981). Known volumes were also filtered and the filter retained to determine chlorophyll-a (Chl-a) concentrations (APHA, 1998) using standard spectrophotometric methods (Perkin Elmer, Lambda 35). 2.3. Sediment sampling In each lake 3 sediment cores were collected using Plexiglass tubes (inner diameter 5 cm, height 30 cm) at the hypolimnetic and littoral sites for sediment characterisation, and 3 or 6 cores (inner diameter 8 cm, height 30 cm) were collected at hypolimnetic and littoral sites respectively, for determination of DO and  dissolved inorganic nitrogen (DIN = NH+4 + NO 2 + NO3 ) fluxes and denitrification rates. Cores were hand collected at the littoral and using a gravity corer at the hypolimnetic sites. Immediately after collection sediment cores were placed in a cool box and returned to the laboratory within 3 h for processing or incubation. Approximately 200 L of site water from each sampling location were collected from ~50 cm above the sediment for core maintenance and laboratory incubations. 2.4. Determination of sediment characteristics In the laboratory, cores were extruded and the upper 0–1 cm sediment depth horizon retained to determine sediment porosity and density, Chl-a and phaeopigment (Pha), organic matter, organic carbon, nitrogen and phosphorous content. Sediment density, porosity and organic matter content, as loss on ignition at 550 °C (LOI) were determined according to Azzoni et al., 2015. Organic carbon and nitrogen contents of dried sediment subsamples were determined using a CHN elemental analyser after removal of carbonates (CHNS-O EA 1108 Carlo Erba) and organic phosphorous was determined following Aspila et al. (1976). Chl-a and Pha contents were determined after extraction of 1 cm3 of sediment with 90% acetone according to Lorenzen (1967). 2.5. Determination of sediment-water column oxygen and DIN fluxes, and denitrification rates In the laboratory cores were submersed with the top open in tanks containing water from the sampling site and maintained in controlled temperature rooms at the temperature measured in the field. Oxic conditions in the tank water collected from the littoral zone was assured by bubbling air with airstones. Hypoxic/ Anoxic conditions in the tank water collected from the hypolimnion were assured by bubbling the water with N2. Cores collected from the littoral stations were maintained under a light/dark cycle close to that of the sampling period, while cores collected from the hypolimnion were kept in the dark. The water inside the cores was gently stirred avoiding sediment resuspension during the preincubation and incubation periods by magnetic stirrers suspended within each core that were driven by a large magnet rotated by an external motor at 40 rpm. Measurements began the day after the sampling. For both flux and denitrification incubations, prior to commencing measurements, the water within the cores was exchanged with the water in the incubation tank. Cores collected in the littoral zone were incubated under both light (n = 3) or dark (n = 3) conditions, in order to detect the effects of autotrophic activity, while cores collected from hypolimnetic sites were incubated only under dark

conditions to mimic the prevailing in situ conditions. Incubation times ranged from 4 (littoral sediments) to 6 (hypolimnetic sediments) hours and were set, based on previous incubations, as the minimum time needed to detect significant changes in solute concentrations and to keep DO concentrations at the end of the incubation within 20% of the initial value (Dalsgaard et al., 2000). Light incubations of littoral sediments were performed under the average irradiance of the sampling period as described in Nizzoli et al. (2014). To initiate flux incubations, the water level in the incubation tanks was lowered to below the core tops and the cores immediately sealed using floating Plexiglass lids. At the beginning of the incubation period before closing the cores and at the end of the incubations, water samples were collected using 50 ml plastic syringes with attached tubing, as preliminary tests demonstrated that changes in solute concentrations where linear over time. Samples for O2 determinations were transferred to glass vials (Exetainers, Labco, High Wycombe, UK), which were filled to overflowing, fixed with Winkler reagents (APHA, 1998) and sealed avoiding trapping  any air bubbles. Samples for NH+4, NO 2 and NO3 determinations were filtered through Whatman GF/F glass fiber filters, transferred to plastic vials, stored at 4 °C and analysed within 24 h, as described for water column samples. Following the flux incubations, the incubation tanks were refilled with site water to re-submerge the cores, which were left uncapped for two hours to re-equilibrate prior to determination of denitrification rates by the Isotope Pairing Technique (Nielsen, 1992). To initiate incubations, the water in the tanks was again lowered to below the core tops, an initial water sample collected and a 15NO 3 solution (98 atom %, Sigma-Aldrich) added to each core to give final concentrations of 2–190 lM 15NO 3 depending on the lake 14NO 3 concentration (final enrichment of 20–60% 15 15 NO N addition to each core was calculated by dif3 ). The actual ference from the water samples collected before and samples collected 10 min after the 15NO 3 addition. The presence of ANAMMOX can interfere with IPT calculations resulting in overestimation of DT because the N2 produced by ANAMMOX cannot be discriminated from that produced by denitrification. Therefore, independence of DT from added 15NO 3 concentration was checked to validate the IPT assumptions and exclude significant overestimation due to ANAMMOX (Risgaard-Petersen et al., 2003). Following the 15NO 3 addition, cores were pre-incubated without lids for 30 min before the start of the incubation to allow the 15  NO 3 tracer to diffuse to the sediment NO3 reduction zone (Nielsen, 1992). Cores were then closed using floating plexiglass lids. Incubation times were the same as those used for the flux incubations and set to ensure that oxygen consumption was less than 20% of the initial O2 concentration, which is a prerequisite for the isotope pairing method (Nielsen, 1992). At the end of the incubations a 7 M ZnCl2 solution was added to each core to a final concentration of 10 mM, in order to prevent further denitrification. Dissolved N2 pools in the water column and sediment porewater were homogenized by gently stirring the sediment with a metal bar. Slurries were briefly allowed to settle and subsamples collected using 50 ml syringes and tubing, transferred to 12 ml Exetainer vials (Exetainer, Labco, High Wycombe, U.K.) filled to overflowing, and poisoned with 200 ml 50% w/v ZnCl2 to prevent further denitrification and the tubes sealed for subsequent analysis of the 29N2 and 30N2 composition of the dissolved N2 pool. 2.6. Calculation of flux, denitrification rates, theoretical pelagic primary production and nitrogen assimilation Oxygen and DIN fluxes were calculated from the difference between final and initial concentration of the target compound

Please cite this article as: D. Nizzoli, D. T. Welsh and P. Viaroli, Denitrification and benthic metabolism in lowland pit lakes: The role of trophic conditions, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134804

D. Nizzoli et al. / Science of the Total Environment xxx (xxxx) xxx

in the water column (Dalsgaard et al., 2000; Nizzoli et al., 2014) according to Eq. (1);

ðCf  CiÞ  V Fx ¼ At

ð1Þ

where, Fx is the flux of species x (mmol m2 h1), Cf and Ci the final and initial concentrations of x (mmol L1), V the volume of the water inside to core (L), t the incubation time (h) and A the sediment surface area inside the core (m2). Total denitrification rates (DT), denitrification of NO 3 diffusing to the sediment from the water column (DW) and denitrification of NO 3 produced within the sediment by nitrification (DN) were calculated from the production rates of 29N2 and 30N2 during the incubations following the equations proposed by Nielsen (1992) as follows:

D15 ¼ p29 þ 2p30 DT ¼ D15  ðp29=2p30Þ
DW ¼ 14NO3 =15NO3  D15 DN ¼ DT  DW 14 where D15 is denitrification of the added 15NO NO 3, 3 the ambient 15  unlabelled NO concentration and NO the isotopically-labelled 3 3 NO 3 concentration at the start of the incubation, and p29 and p30 are the production rates of 29N2 and 30N2, respectively. The abundance of 29N2 and 30N2 in the dissolved N2 pool was determined by mass spectrometry at the National Environmental Research Agency, Silkeborg, Denmark, as previously described by Risgaard-Petersen and Rysgaard (1995). Net daily sediment-water column fluxes and denitrification rates were calculated from the measured light and dark rates multiplied by the average light and dark periods at the time of sampling for the littoral stations and as 24 times the dark rates for hypolimnetic sediments. The trophic state of the lakes was assessed on the basis of the mean volume weighted concentration of Chl-a and TP, and according to the fixed boundaries of OECD standards (Vollenweider and Kerekes, 1982). Benthic metabolism was assessed at the littoral stations in each lake using the Benthic Trophic State Index – BTSI (Rizzo et al., 1996). BTSI is based on the net oxygen metabolism resulting from gross production (net production in light + dark respiration) and dark community respiration, and allows classification of the considered systems into four categories: (0) fully heterotrophic, (1) net heterotrophic, (2) net autotrophic, and (3) highly autotrophic. The theoretical nitrogen demand for growth of benthic microalgae was estimated from net benthic production (NBP) calculated from the daily net oxygen fluxes and converted into carbon equivalents assuming a photosynthetic quotient of 1.2 (Schramm et al., 1984; Wetzel and Likens, 1991). The nitrogen demand to support this production were calculated assuming a molar C:N ratio of 9 for benthic microalgae (Sundbäck et al., 2000). Pelagic phytoplankton primary production (mg C m3 d1) was estimated using the published relationship for pelagic metabolism (Weilhartner et al., 2012; Wetzel, 2001) using the in situ Chl-a data, as follows:

Pelagic production ¼ chlorophyll a  22:9  42:6 Theoretical pelagic Nr assimilation was calculated from the estimated pelagic primary production assuming a photosynthetic quotient of 1.2 and a molar C:N ratio of 6.62 for phytoplankton (Schramm et al., 1984; Wetzel and Likens, 1991).

5

2.7. Statistical analyses of data Flux data and denitrification rates were analysed with two way generalized least square (GLS) models. A first model was run with lake and depth (hypolimnetic sediment, littoral sediment in dark) as the main fixed factors (interaction included) to test the effect of depth and lake. A second model was run on data from the littoral sites only with lake and condition (dark and light) as the main fixed factors (interaction included) to test the effect of the activity of primary producers on fluxes and process rates, as the deep hyplimnetic sites are in constant darkness and were therefore only incubated under dark conditions. To deal with observed heteroscedasticity, we used the argument ‘‘weights” within the function gls(), and the function varIdent() to specify variance models. An a posteriori comparison of the means was performed using a post hoc Tukey test. Statistical analyses were conducted using the nmle (Pinheiro et al., 2016) and emmeans (Lenth et al., 2018) packages in the R software v. 3.5.1 (R Core Team, 2018). Relationships between environmental factors (sediment LOI, Chl-a, and Pha contents, and DO and NO 3 concentrations in the water above the sediment), and sediment oxygen consumption and Nr process rates (denitrification rates, NH+4 and NO 3 fluxes) were analysed using Spearman correlation analysis in the R software v. 3.5.1 (R Core Team, 2018). All results are presented as mean ± standard error. 3. Results 3.1. Water column physico-chemical characteristics Chl-a and total phosphorus concentrations in surface waters varied among lakes (Table 2). Chl-a concentrations measured in the surface waters of IG and BL (mean 10.9 mg L1) were approximately 3-fold higher than those measured in the other lakes (mean 3.5 mg L1). Total phosphorus concentrations were also 2.8-fold higher in more eutrophic BL and IG (mean 1.45 mM) compared to the other lakes (mean 0.51 mM). Two of the studied lakes (CM and CS) can be classified as oligo-mesotrophic, two (IG and BL) as eutrophic, while LV was mesotrophic. All lakes were thermally stratified with water temperatures ranging from 26 to 28 °C in the surface waters of the hypolimnetic and littoral sites respectively (Fig. 1S in Supplementary material) to ~13 °C in the bottom waters of all lakes, except CS where the bottom water was ~8 °C. DO concentrations were close to or above saturation in the surface waters of all lakes, while waters were hypoxic (CM and CS) to less than 1 mg L1 (LV, IG and BL) below the thermocline at the hypolimnetic sites (Fig. 1S in Supplementary material). NO 3 was the dominant form of DIN in the surficial water at both littoral and hypolimnetic sites in all the lakes accounting 80% or more of total DIN (Fig. 1S in Supplementary material). However, actual surface water NO 3 concentrations varied substantially between lakes ranging from 56 ± 3 mM (BL) to 451 ± 2 mM (CM). Although there was a tendency for NO 3 concentrations to decrease with water depth, especially in the more eutrophic lakes, NO 3 was present in the bottom waters above the sediment of all lakes, at concentrations ranging from ~4 mM in hypertrophic BL, to between 28 and 40 mM in the mesotrophic and hypereutrophic lakes (LV, IG), and at 216 and 508 mM in oligotrophic CM and CS, respectively. NH+4 concentrations were typically <10 mM in surface and deep waters in CM and CS, and in the surface waters of the other lakes (Fig. 1S in Supplementary material). However, in BL, LV and IG, NH+4 concentration increased with water depth, attaining 50– 80 mM in the hypolimnion.

Please cite this article as: D. Nizzoli, D. T. Welsh and P. Viaroli, Denitrification and benthic metabolism in lowland pit lakes: The role of trophic conditions, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134804

6

D. Nizzoli et al. / Science of the Total Environment xxx (xxxx) xxx

Table 2 Water column and sediment physicochemical characteristics and the benthic trophic state classification (BTSI) of the five investigated lakes (CM = Ca’ Morta, CS = Ca’ Stanga, LV = Lago Verde, IG = Isola Giarola, BL = Bosco della Lite) at littoral (L) and hypolimnetic (H) sites. Units are mg L1 chlorophyll a (chl-a), mM for total phosphorus (TP), ammonium 1 (NH+4) and nitrate (NO wet sediment for sediment 3 ), % of dry weight for organic matter (LOI), organic carbon (C), organic nitrogen (N) and organic phosphorous (P), and mg ml cholophyll a (chl-a) and phaeopigments (Pha). All the data are mean ± standard error (standard error in parenthesis). BTSI was only calculated for the littoral sites (n.a. = not applicable). CM

CS

LV

IG

BL

L

H

L

H

L

H

L

H

L

H

Water TP Chl-a NH+4 NO 3

0.4 (0.0) 2.1 (0.1) 1.6 (0.3) 451.5 (1.8)

0.7 (0.0) 2.9 (0.8) 7.9 (0.5) 477.8 (5.8)

0.6 (0.1) 4.7 (0.0) 0.9 (0.1) 243.7 (10.6)

0.8 (0.3) 3.0 (0.2) 6.9 (0.8) 269.1 (2.6)

0.5 (0.0) 3.7 (0.1) 1.9 (0.1) 156 (0.3)

0.7 (0.2) 6.1 (1.0) 12.3 (6.7) 127.7 (19.4)

1.4 (0.5) 7.1 (1.8) 8.0 (6.0) 51.1 (18.3)

2.3 (0.5) 7.9 (1.2) 45.4 (3.0) 37.9 (2.0)

1.5 (0.4) 14.7 (3.3) 3.4 (1.6) 56.4 (2.9)

1.7 (0.7) 3.3 (0.4) 64.7 (12.6) 3.4 (0.5)

Sediment LOI Chl-a Pha C N P BTSI

2.12 (0.12) 15.79 (2.42) 7.56 (0.90) 0.67 (0.04) 0.08 (0.00) 0.03 (0.00) 3

3.20 0.55 1.11 0.71 0.08 0.02 n.a.

2.51 5.43 4.28 0.79 0.08 0.03 3

3.67 (0.20) 2.17 (0.19) 6.33 (0.33) 0.85 (0.04) 0.1 (0.00) 0.01 (0.00) n.a.

2.01 (0.20) 14.24 (2.53) 6.53 (0.14) 0.83 (0.10) 0.08 (0.01) 0.02 (0.01) 2

6.19 (0.64) 9.19 (5.00) 21.45 (6.15) 2.76 (0.02) 0.28 (0.00) 0.01 (0.01) n.a.

3.23 4.64 7.65 1.34 0.11 0.02 1

3.74 (0.18) 5.41 (0.3) 12.93 (0.60) 1.24 (0.01) 0.14 (0.00) 0.01 (0.01) n.a.

3.11 2.37 4.27 1.15 0.09 0.01 1

3.29 4.14 9.94 1.33 0.11 0.01 n.a.

(0.07) (0.02) (0.17) (0.01) (0.00) (0.01)

(0.15) (0.74) (0.86) (0.02) (0.00) (0.00)

(0.34) (0.48) (0.66) (0.01) (0.00) (0.01)

(0.31) (0.16) (0.60) (0.04) (0.01) (0.01)

(0.07) (0.88) (2.27) (0.20) (0.00) (0.01)

3.2. Sediment physico-chemical characteristics LOI was similar among lakes (2.0–3.8%), with the exception of the hypolimnetic sediment in LV (6.2%) and generally LOI was higher in the hypolimnetic compared to littoral sediments (Table 2). These differences were also reflected in the sediment elemental organic C and N contents which showed a similar variation across the 5 lakes and were approximately two-fold higher in hypolimnetic compared to littoral sediments. Sediment organic P contents were also similar across lakes, but tended to be lower at hypolimnetic compared to littoral sites (Table 2). Consequently, whilst C:N ratios of the sediment organic matter pools were similar in littoral and hypolimnetic sediments, C:P and N:P ratios were substantially higher in the hypolimnetic compared to littoral sediments (data not shown). Sediment total photosynthetic pigment (Chl-a + Pha) content was variable between lakes ranging from 1.6 to 30 mg ml1. Sediment Chl-a contents, which are a proxy for active microalgal biomass, and Pha contents, which are a proxy for degraded microalgal biomass, were equally variable, ranging between 0.6 and 15.8, and 1.1 and 24.5 mg ml1, respectively. Chl-a concentrations however, tended to be higher in littoral compared to hypolimnetic sediments, while Pha concentration tended to be higher in hypolimnetic sediments. The ratio between Chl-a and Pha concentrations was variable between lakes and sites, ranging between 0.3 and 2.1, with the highest ratios measured at the littoral sites in CM, CS and LV and the lowest in hypolimnetic sediments.

3.3. Benthic oxygen metabolism Under dark conditions sediment O2 fluxes differed significantly between the hypolimnetic and littoral sediments and among lakes (Fig. 2 and Table 3). On average sediment oxygen uptake was 3.4 times higher in littoral (2001 ± 138 mmol O2 m2 h1) compared to hypolimnetic sites (575 ± 47 mmol O2 m2 h1). Fluxes were significantly higher in more eutrophic IG and BL (1542 ± 448 and 1472 ± 387 mmol O2 m2 h1, respectively) than in the oligotrophic CM (803 ± 295 mmol O2 m2 h1) (Fig. 2). Oxygen fluxes in hypolimnetic sites exhibited a significant negative correlation with sediment LOI and pigment contents and a positive correlation with NO 3 concentration (Table 4), while they exhibited a significant positive correlation with sediment Chl-a content and NO 3 concentration at littoral sites (Table 4).

Fig. 2. Upper panel: benthic oxygen fluxes measured at hypolimnetic (dark only) and littoral (light and dark) sites in each lake. Letters and symbols indicate the results of the post hoc Tukey test. Letters indicate similarities (e.g AA or aa) or significant differences (e,g, AB or ab) amongst sites, with capital letters below the xaxis indicating differences among hypolimnetic sites (model 1) and lowercase letters differences among littoral sites under light conditions (model 2); a * indicates a significant difference between light and dark fluxes at the littoral site of that lake (model 2). p < 0.05 in all cases. Lower panel: gross primary production measured at littoral sites (bars) and estimated pelagic gross primary production (line). Error bars represent standard error (n = 3).

At littoral sites benthic oxygen fluxes differed with light conditions and lakes, but there was a significant interaction between light conditions and lake (Table 3). In eutrophic IG and BL the ben-

Please cite this article as: D. Nizzoli, D. T. Welsh and P. Viaroli, Denitrification and benthic metabolism in lowland pit lakes: The role of trophic conditions, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134804

7

D. Nizzoli et al. / Science of the Total Environment xxx (xxxx) xxx

Table 3  Summary of the results of two way generalized least square (GLS) models of benthic oxygen consumption (O2), ammonium (NH+4), nitrite (NO 2 ), nitrate (NO3 ) and dissolved inorganic nitrogen (DIN) fluxes, total denitrification (DT), denitrification of nitrate diffusing from the water column (DW) and coupled nitrification–denitrification (DN). Model 1 was run with lake and depth (hypolimnetic and littoral sites) as main fixed factors using data from the dark incubations. Model 2 was run only with data from littoral sites, with lake and light condition (dark and light) as main fixed factors. Significant (p < 0.05) outcomes are highlighted in bold. Souce of variation

O2 P

DT

DW

DN

NH4

NO2

NO3

DIN

DF Model 1 Lake Depth Lake*Depth

4 1 4

<0.001 <0.001 0.411

<0.001 <0.001 0.009

<0.001 <0.001 <0.001

<0.001 <0.001 0.06

<0.001 <0.001 0.021

0.028 <0.001 0.037

<0.001 <0.001 <0.001

<0.001 <0.001 <0.001

Model 2 Lake Condition Lake*Light

4 1 4

<0.001 <0.001 <0.001

<0.001 0.100 0.741

0.011 <0.001 0.405

0.090 0.063 0.350

0.004 0.002 0.835

<0.001 <0.001 0.238

<0.001 0.006 0.058

<0.001 0.041 0.102

thos was a net oxygen sink in the light (average 1874 ± 271 mmol O2 m2 h1) and fluxes during light incubations were similar to those measured under dark conditions (average 2376 ± 219 mmol O2 m2 h1) (Fig. 2). Conversely, in CM, CS and LV oxygen fluxes differed significantly with light conditions. Net oxygen production was measured for littoral sediments in CM, CS and LV under light conditions, with significantly higher production measured in CM and CS (average 4434 ± 577 mmol O2 m2 h1) compared to LV (average 1278 ± 211 mmol O2 m2 h1). Benthic oxygen fluxes measured under light conditions at the littoral sites were positively correlated with sediment Chl-a and NO 3 concentration, and negatively correlated with sediment LOI (Table 4). Rates of gross primary production at littoral sites (calculated from differences between mean light and dark oxygen fluxes) were highly variable between lakes and ranged from 3 to 62 mg C m2 h1, with the highest production occurring in CM and CS and lowest in IG and BL (Fig. 2). Based on the measured oxygen fluxes, according to the Benthic Trophic State Index, the status of the littoral sites in each lake can

be classified as highly autotrophic for CM and CS (class 3), net autotrophic for LV (class 2), and net to fully heterotrophic for IG and BL (class 1–0). 3.4. Sediment-water column dissolved inorganic nitrogen fluxes Under dark conditions NO 3 fluxes were always directed toward the sediment and there was a significant strong interaction between lakes and depths (Table 3; Fig. 3). Under dark conditions sediment NO 3 consumption was on average two times greater for littoral (256 ± 27 mmol m2 h1) compared to hypolimnetic sediments (119 ± 14 mmol m2 h1), but differences in fluxes between littoral and hypolimnetic sites were only statistically significant for CM and CS (Fig. 3). Hypolimnetic sediment NO 3 consumption was highest in LV and CS (average 176 ± 14 mmol m2 h1) and lowest in CM and BL (67 ± 13 mmol m2 h1). Sediment-water column NO 3 fluxes in the littoral zone differed significantly between lakes and light conditions (Table 3). NO 3

Table 4 Correlations (Spearman’s rho) between environmental factors (sediment organic matter (LOI); sediment chlorophyll (Chl a) and phaeopigment (Pha) concentration; water column +  oxygen (O2) and nitrate (NO 3 ) concentration above the sediment; sediment water column oxygen (O2 flux), ammonium (NH4 flux) fluxes and nitrate consumption (NO3 cons); total denitrification (DT), denitrification of water column nitrate (DW) and denitrification coupled to nitrification (DN) rates. Significant (p < 0.05) correlations are highlighted in bold.

Hypolimnion

Littoral light

Littoral dark

LOI

chl a

Pha

O2

NO 3

O2 flux

NH4 flux

NO 3 cons

DT

b-chl Pha O2 NO 3 O2 flux NH4 flux NO3 cons DT DW

0.90 0.90 0.30 0.30 0.76 0.38 0.65 0.89 0.93

1.00 0.10 0.60 0.85 0.64 0.33 0.69 0.78

0.10 0.60 0.85 0.64 0.33 0.69 0.78

0.69 0.11 0.70 0.84 0.57 0.43

0.61 0.75 0.38 0.06 0.13

0.61 0.25 0.53 0.62

0.31 0.06 0.26

0.77 0.71

0.96

b-chl Pha O2 NO 3 O2 flux NH4 flux NO3 cons DT DW DN

0.80 0.10 0.00 0.60 0.63 0.27 0.17 0.05 0.04 0.28

0.40 0.30 0.90 0.74 0.19 0.64 0.15 0.02 0.09

0.89 0.30 0.08 0.43 0.43 0.12 0.04 0.32

0.10 0.22 0.51 0.11 0.03 0.06 0.06

0.87 0.32 0.85 0.38 0.16 0.24

0.66 0.63 0.15 0.01 0.18

0.07 0.31 0.36 0.19

0.51 0.19 0.45

0.74 0.26

0.35

b-chl Pha O2 NO 3 O2 flux NH4 flux NO3 cons DT DW DN

0.80 0.10 0.00 0.60 0.43 0.31 0.38 0.44 0.45 0.35

0.40 0.30 0.90 0.69 0.38 0.70 0.48 0.47 0.31

0.89 0.30 0.28 0.19 0.28 0.00 0.02 0.14

0.09 0.07 0.48 0.13 0.09 0.16 0.10

0.74 0.46 0.87 0.61 0.58 0.42

0.64 0.43 0.42 0.26 0.19

0.12 0.06 0.17 0.13

0.70 0.73 0.41

0.94 0.39

0.59

DW

Please cite this article as: D. Nizzoli, D. T. Welsh and P. Viaroli, Denitrification and benthic metabolism in lowland pit lakes: The role of trophic conditions, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134804

8

D. Nizzoli et al. / Science of the Total Environment xxx (xxxx) xxx

 Fig. 3. Ammonium (NH+4), nitrite (NO 2 ), nitrate (NO3 ) and dissolved inorganic nitrogen (DIN) fluxes measured at the hypolimnetic and littoral sites of each lake. Error bars represent standard errors (n = 3). Letters and symbols indicate the results of the post hoc Tukey test. Letters indicate similarities (e.g AA or aa) or significant differences (e.g. AB or ab) amongst sites, with capital letters below the x-axis indicating differences among hypolimnetic sites (model 1) and lowercase letters above the x-axis differences among littoral sites (model 2); a # indicates a significant difference in dark fluxes between the hypolimnetic and littoral sites of that lake. p < 0.05 in all cases.

fluxes were always directed toward the sediment during both light and dark incubations, however light exposure significantly reduced sediment NO 3 consumption, with mean fluxes measured under light conditions (190 ± 35 mmol m2 h1) being on average 40% lower than those measured under dark conditions (–323 ± 51 mmol m2 h1). Among the littoral sites, sediment NO 3 consumption was significantly greater in oligotrophic CM and CS (average 440 ± 35 mmol m2 h1), whereas NO 3 consumption was significantly lower in eutrophic BL (–33 ± 22 mmol m2 h1), (Fig. 3). At the hypolimnetic sites NO 3 consumption exhibited a significant positive correlation with sediment LOI and water column DO (Table 4). Whereas, at the littoral sites NO 3 consumption exhibited a significant positive correlation with sediment Chl-a content, water column DO and oxygen fluxes under light conditions, and was positively correlated with sediment Chl-a content and water column NO 3 concentration under dark conditions. Under dark conditions sediment-water column NH+4 fluxes differed amongst lakes and between littoral and hypolimnetic sites, with a significant interaction between these factors (Table 3). NH+4 fluxed from the sediment to the water column, with the exception of the hypolimnetic sites in oligotrophic CM and CS, which exhibited low rates of NH+4 uptake (Fig. 3). NH+4 effluxes were typically higher at the littoral (45 ± 8 mmol m2 h1) compared to the hypolimnetic sites (3 ± 1 mmol m2 h1). This difference was, however, only statistically significant for CM, CS and LV (Fig. 3). At the hypolimnetic sites NH+4 fluxes exhibited significant positive correlations with sediment pigment contents, and a negative cor-

relation with water column DO and NO 3 concentrations, and oxygen fluxes (Table 4). At littoral sites NH+4 fluxes differed significantly with light conditions and among lakes (Table 3). Light exposure either reduced the NH+4 efflux observed under dark conditions (LV, IG and BL) or reversed the flux with the sediments exhibiting low uptake of NH+4 during light incubations (CM and CS). Overall, across all lakes NH+4 fluxes measured under light conditions (29 ± 16 mmol m2 h1) were ~70% lower than those determined in the dark (99 ± 16 mmol m2 h1). Highest NH+4 efflux was measured in IG (142 ± 24 mmol m2 h1), while the lowest effluxes were measured in CM and CS (26 ± 20 mmol m2 h1), where the effluxes measured under dark conditions were partially offset by low rates of sediment uptake in the light (Fig. 3). NH+4 fluxes measured under light conditions at the littoral sites exhibited a significant negative correlation with benthic oxygen fluxes (Table 4). Whereas, NH+4 fluxes measured under dark conditions only exhibited a significant negative correlation with benthic oxygen fluxes. Typically, NO 2 fluxed to the water column and the magnitude of efflux was greater for littoral (mean 45 ± 8 mmol m2 h1) compared to hypolimnetic (mean 3 ± 1 mmol m2 h1) sites, but this depth difference was only significant for CM, CS and IG (Fig. 3). At the littoral sites NO 2 fluxes differed significantly amongst lakes, and with light conditions (Table 3). Light exposure significantly reduced NO 2 efflux from the sediment, with effluxes measured in light (23 ± 6 mmol m2 h1) being on average ~50% of those determined during dark incubations (45 ± 8 mmol m2 h1).

Please cite this article as: D. Nizzoli, D. T. Welsh and P. Viaroli, Denitrification and benthic metabolism in lowland pit lakes: The role of trophic conditions, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134804

D. Nizzoli et al. / Science of the Total Environment xxx (xxxx) xxx

9

tions were on average higher at the littoral (158 ± 33 mmol m2 h1) compared to the hypolimnnetic (87 ± 26 mmol m2 h1) sites, but differences were only significant in CM and CS. The benthos was a net sink for DIN in CM, CS and LV at both hypolimnetic (168 ± 15) and littoral (-312 ± 47 mmol m2 h1) sites (Fig. 3). In contrast in eutrophic IG and BL the sediment was a net source of DIN to the overlying water at both littoral (35 ± 9) and hypolimnetic (22 ± 20 mmol m2 h1) sites. In the littoral zone DIN fluxes differed significantly amongst lakes and with light conditions (Table 3). Light exposure reduced DIN consumption, with fluxes measured in light (138 ± 39 mmol m2 h1) being on average the 20% lower than those in the dark (-178 ± 48 mmol m2 h1).

3.5. Sediment denitrification rates

Fig. 4. Rates of total denitrification (DT), denitrification of water column nitrate (DW) and denitrification coupled to nitrification (DN), measured at the hypolimnetic and littoral sites of each lake. Error bars represent standard error (n = 3). Letters and symbols indicate the results of the post hoc Tukey tests. Letters indicate similarities (e.g AA or aa) or significant differences (e,g, AB or ab) amongst sites, with capital letters below the x-axis indicating differences among hypolimnetic sites (model 1) and lowercase letters differences among littoral sites (model 2); a # indicates a significant difference in dark rates between hypolimnetic and littoral sites in that lake. p < 0.05 in all cases.

Overall benthic fluxes of DIN were largely driven by the NO 3 fluxes and varied from a minimum of 500 to a maximum of 74 mmol m2 h1 (Fig. 3). DIN fluxes measured under dark condi-

Sediment total denitrification rates measured under dark conditions were higher (Table 3, Fig. 4) at littoral (181 ± 15 mmol m2 h1) compared to hypolimnetic sites (54 ± 12 mmol m2 h1). The degree of difference between littoral and hypolimnetic rates however, was not consistent across the lakes, ranging from 235-fold (CM) to 1.9-fold (LV). Hypolimnetic DT rates in eutrophic CM were below 10 mmol m2 h1 and significantly lower compared to all other lakes, but rates were also significantly lower in oligotrophic BL than at all other sites, except CM (Fig. 4). Whereas, highest hypolimnetic rates of DT were measured in mesotrophic LV (116 ± 7 mmol m2 h1). Total denitrification rates in hypolimnetic sites exhibited significant positive correlations with sediment LOI, pigment content, DO concentration and NO 3 consumption, and negative correlations with oxygen fluxes (Table 4), but were not significantly correlated with water column NO 3 concentration (Table 4; Fig. 5). At the littoral sites, DT varied significantly between lakes, but not with light conditions (Table 3), although rates under light conditions were on average 22% lower compared to those under dark conditions. Highest DT rates under both light and dark conditions were measured in CS (mean 291 ± 18 mmol m2 h1) and were significantly higher than those in other lakes, which were similar (Fig. 4). DT rates in littoral sediments were positively correlated  with water column NO 3 concentration and benthic NO3 consumption (Table 4; Fig. 5) under dark conditions, but no correlations existed between DT and any of the measured environmental variables under light conditions (Table 4). Sediment DT rates were mainly sustained by direct reduction of NO 3 diffusing from the overlying water (DW) with average rates across lakes, sites and conditions of 100 ± 11 mmol m2 h1, which were approximately 4-fold higher than those of denitrification coupled to nitrification within the sediment (DN) (25 ± 5 mmol m2 h1). In hypolimentic sediments water column  NO 3 was almost the exclusive source of NO3 for denitrification with DW accounting for on average 90% of DT. DW rates measured under dark condition were ~3.5-fold higher at the littoral compared to hypolimnetic sites, but there was a significant interaction between lake and depth conditions (Table 3) and consequently this difference between shallow and deep sampling sites varied considerably across the lakes (Fig. 4). The greatest differences between littoral and hypolimnetic DW rates occurred in the oligotrophic CM and eutrophic BL lakes. In both these lakes, hypolimnetic DW rates were >10 mmol m2 h1 and significantly lower than in all other lakes. Highest hypolimnetic DW rates were measured in mesotrophic LV (111 ± 11 mmol m2 h1), where rates were significantly greater than in all other lakes (Fig. 4). DW at the hypolimnetic sites was positively correlated with sediment LOI, Pha content and NO 3 consumption, and negatively correlated with oxygen fluxes (Table 4).

Please cite this article as: D. Nizzoli, D. T. Welsh and P. Viaroli, Denitrification and benthic metabolism in lowland pit lakes: The role of trophic conditions, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134804

10

D. Nizzoli et al. / Science of the Total Environment xxx (xxxx) xxx

Fig. 5. Relationship between total denitrification rates and water column NO 3 concentration for hypolimnetic (left panel) and littoral (right panel) sites. Data from light incubations are shown as open circles and from dark incubations as solid black circles, two letter codes indicate the study lake and error indicate the standard error (n = 3).

At the littoral sites DW differed significantly between lakes and with light conditions (Table 3). Across all lakes DW rates measured during light incubations (mean 118 ± 17 mmol m2 h1) were on average ~30% lower compared to those under dark conditions (164 ± 18 mmol m2 h1), however, the degree of difference varied considerably between lakes, being greatest in the more oligotrophic and marginal in eutrophic IG and BL (Fig. 4). At littoral sites, DW was positively correlated with water column NO 3 concentration and NO 3 consumption under dark conditions (Table 4). In hypolimnetic sediments rates of coupled nitrification–denitri fication (DN) were only measurable in CS (28 ± 7) and LV (7 ± 3 mmol m2 h1). At the littoral sites, DN showed no significant differences with either lake or light conditions (Table 3), and average measured rates across lakes were similar under light (38 ± 13) and dark (33 ± 6 mmol m2 h1) conditions. There was however, a tendency for higher light compared to dark rates in the more oligotrophic lakes (CM and CS), whereas, light and dark rates were very similar or higher under dark conditions in the eutrophic IG and BL (Fig. 4). DN was negatively correlated with DW under dark, but not light conditions (Table 4). 4. Discussion In this study we analysed the capacity of the benthic system of pit lakes to remove Nr loads and how the fate of Nr varied with lake trophic state and water depth. The studied lakes differed considerably in terms of their trophic conditions and water column NO 3 concentrations, as has been observed in other pit lakes in the Po river floodplain and other watersheds (Kattner et al., 2000; Tavernini et al., 2009), although this is not always the case (Søndergaard et al., 2018). Across this gradient of trophic conditions, the benthic system of the lakes was, on average, a net sink for water column DIN and thus the lake sediments contributed to reducing nitrogen loads. However, benthic Nr process rates and the fate of the Nr were highly variable both within and amongst the lakes. 4.1. Denitrification as a relevant benthic process in pit lakes Denitrification is a relevant benthic Nr process in pit lakes that can potentially mitigate Nr loads. Measured denitrification rates in

the five studied pit lakes were within the upper range of values in marine environments (Pina-Ochoa and Álvarez-Cobelas, 2006) and comparable to those of natural and manmade freshwater aquatic lakes and ponds (e.g. Helmer and Labroue, 1993; Scott et al., 2008; Herrman and White, 2008; Bruesewitz et al., 2011; Nizzoli et al., 2018). Despite the relatively recent origin of the studied lakes, the sediments consumed on average from ~2 (hypolimnion) to ~4 (littoral zone) mmol m2 d1 of DIN. Comparing these DIN consumption rates to measured N2 production provides insights on the relevance of denitrification to the benthic Nr cycle. On average, N2 production via denitrification accounted for 60 and 96%, respectively of the total DIN consumed by the littoral and hypolimnetic sediments. These finding highlight the importance of denitrification as a benthic Nr elimination process in the investigated lakes. Benthic denitrification rates are driven by two sources of NO 3, pre-existing NO 3 diffusing into the sediment from the overlying water and that produced by nitrification within the sediment (Nielsen, 1992). Differentiating between these NO 3 sources can provide useful insights into the function of the benthos (Gardner and McCarthy, 2009). As denitrification coupled to nitrification (DN) represents the capacity of the system to dissipate internal nitrogen loads during the remineralisation of organic nitrogen, whereas direct denitrification of water column NO 3 (DW) represents the capacity of the sediment to eliminate external NO 3 loads from groundwater inputs and/or surface runoff. In the studied pit lakes, benthic denitrification was primarily sustained by reduction of NO 3 diffusing from the overlying water, with measured rates of DW ranging between 1.1 and 3.4 mmol m2 d1. Overall, DW accounted for 73–100% of DT across all sites, confirming that water column NO 3 was the primary driver for N2 production at both littoral and hypolimnetic sites in all 5 lakes. These results are in broad agreement with previous studies demonstrating that denitrification is strongly influenced by NO 3 availability in the overlying water (Rissanen et al., 2013), and that lakes located in watersheds with a high proportion of agriculture have the highest denitrification potential (Bruesewitz et al., 2011). The lakes investigated in this study are located in an area heavily exploited for crop and livestock production (Viaroli et al., 2018) and the excess Nr resulting from fertilizer and manure application can be leached to ground and surface waters at high rates (Sutton et al., 2011; Billen et al., 2013). Consequently, NO 3 leaching from surrounding

Please cite this article as: D. Nizzoli, D. T. Welsh and P. Viaroli, Denitrification and benthic metabolism in lowland pit lakes: The role of trophic conditions, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134804

D. Nizzoli et al. / Science of the Total Environment xxx (xxxx) xxx

farmlands can explain the high NO 3 concentrations in the studied lakes and this in turn would drive the high denitrification rates measured in the lake sediments. However, denitrification rates were only weakly correlated with water column NO 3 concentration at the littoral sites and exhibited no significant correlation with NO 3 concentration at the hypolimnetic sites, or littoral sites under light conditions (Fig. 5). Therefore, whilst NO 3 availability is a prerequisite for denitrification, variations in water column NO 3 concentration alone cannot explain the observed variability of denitrification rates in the studied lakes, indicating that other factors are also important. 4.2. Factors controlling intra and inter lake variability in denitrification rates Benthic metabolism and nitrogen process rates varied spatially both within and among lakes as previously reported (e.g. Saunders and Kalff, 2001; McCrackin and Elser, 2010; Rissanen et al., 2013). In this study, denitrification rates, along with DIN fluxes and benthic (dark) community respiration rates, were all significantly greater for littoral compared to hypolimnetic sites, indicating that the littoral zone represents a metabolic ‘‘hot spot” within the lakes (den Heyer and Kalff, 1998; Wetzel, 2001; Bruesewitz et al., 2011; Nizzoli et al., 2018). With the exception of BL, NO 3 concentrations were similar in deep and littoral waters in the studied lakes and therefore differences in NO 3 availability alone cannot explain the significant differences observed in within lake denitrification rates between littoral and hypolimnetic sites. Previous studies have indicated that in addition to NO 3 availability, denitrification rates are also influenced by factors such as, temperature, sediment organic matter enrichment and primary production (Dunn et al., 2012; Nizzoli et al., 2014; Pina-Ochoa and Álvarez-Cobelas, 2006; Sundbäck et al., 2000). Denitrification rates, as with other microbial rates, are highly sensitive to temperature and relatively small changes in temperature can result in large changes in bacterial growth and metabolic rates (White et al., 1991; den Heyer and Kalff, 1998; Uusheimo et al., 2018). This relationship has been standardised using the Q10 temperature coefficient value (Verstraete and Focht, 1977). All the investigated lakes were thermally stratified at the time of sampling and exhibited similar thermal profiles with the thermocline occurring at between approx. 4 and 8 m depth. Due to this summer stratification the temperature of epilimnetic water overlaying the littoral sites was from 11 to 18 °C higher compared to the hypolimnion. Based on the Q10 value of 2 for denitrifying bacteria (Verstraete and Focht, 1977), the measured temperature differences could explain differences in denitrification rates of 2–3.5 fold between littoral and hypolimnetic sites in the studied lakes. Measured denitrification rates differed by ~2 (CS, LV and IG), 23 (BL) and 235-fold (CM) between littoral and hypolimnetic sites. Therefore, while temperature difference between the surficial and deep waters can account for the greater denitrification in the littoral zones of CS, LV and IG, it cannot explain the large differences in rates between littoral and hypolimnetic sites in CM and BL, and consequently, other factors must be involved. The quantity and quality (lability) of sediment organic matter pools influence denitrification both directly as electron donor for respiration processes, including denitrification, and indirectly, by increasing aerobic respiration rates and therefore reducing oxygen penetration depth into the sediment (Rysgaard et al., 1994; Sloth et al., 1995; Small et al., 2014). This decreases the diffusional path length of NO 3 from the overlying water to the denitrification zone, which increases the NO 3 flux and therefore potential denitrification rates (Revsbech et al., 2005). Organic enrichment also favours oxygen consumption and depletion in the hypolimnion, which in turn can negatively influence denitrification sustained by NO 3 pro-

11

duction by nitrification within the sediment. At low oxygen concentration nitrification is impaired, which precludes the supply of NO 3 to denitrifiers (Rysgaard et al., 1994). Inter lake variation of hypolimnetic denitrification and sediment oxygen consumption confirm this dependency as denitrification rates were positively correlated with sediment organic matter and pigments concentrations, and with sediment oxygen consumption (Table 4). Therefore, in addition to temperature low organic matter content can limit denitrification rates. However, in the most oligotrophic lake (CM), LOI was lower in littoral compared to deep sediments and therefore its availability cannot explain the lower rates measured in the deep area of the lake. Yet, the sediment Chl-a to Pha ratio was higher in littoral (2.17) compared to deep sediments (0.51). This difference suggests that sediment organic matter in the hypolimnion was richer in degraded microalgal biomass which would be indicative of lower organic matter quality. In this lake the negative NH+4 flux also indicates Nr limitation and therefore bacterial assimilation of water column Nr, which is also indicative of low quality (high C:N ratio) organic matter sources (Canfield et al., 2005), and additionally may also induce competition for NO 3 between denitrification and microbial Nr assimilation for growth. Thus, in CM differences in the sediment organic matter pools and potential competition between denitrification and Nr assimilation may explain rates of denitrification differing by much more than would be predicted by the temperature differences. On the other hand, in the eutrophic BL differences in organic matter quality cannot explain the large difference in denitrification rates between the littoral and hypolimnetic sites, as Chl-a:Pha ratios were similar (~0.5) and both sediments were equally large sources of NH+4 to the water column, indicating that Nregeneration rates exceeded the growth requirements of the microbial community (Canfield et al., 2005). In this lake however, the difference in denitrification rates between littoral and deep sediments (108 mmol N m2 h1) was very much greater than the 2 1 difference in NO h ) and benthic NO 3 fluxes (9 mmol N m 3 consumption at the hypolimnetic site exceeded DW rates by ~8-fold, suggesting that other nitrate consuming processes occurred in hypolimnetic sediment. In the absence of significant rates of microbial NO 3 assimilation, DNRA is the most plausible sink for  this excess NO 3 . DNRA competes with denitrification for NO3 , which is the electron acceptor for both processes (Welsh et al., 2001) and can outcompete denitrification for NO 3 under eutrophic conditions, as it is favoured by high electron donor (organic matter) to electron acceptor (NO 3 ) ratios (Burgin and Hamilton, 2007; Azzoni et al., 2015; Hardison et al., 2015). Unlike denitrification, which mitigates eutrophication by eliminating NO 3 as gas+ eous end-products, DNRA traps and recycles NO 3 as NH4 and thereby enhances and reinforces eutrophication processes. Although, NO 3 in the overlying water was the dominant source of NO 3 fuelling denitrification in the studied sediments and DW accounted for the bulk of DT, significant rates of DN of up to 234 mmol m2 h1 were measured in the studied lakes. Rates of DN measured in the hypolimnion were significantly lower than those determined in littoral sediments and negatively correlated with DO concentration, indicating that oxygen availability in the surficial sediments limited rates of nitrification and hence DN. This would be expected, as NH+4, the electron donor for nitrification, typically effluxed from the sediments, indicating that it was abundant, and under such conditions nitrification rates are constrained by the availability of oxygen, the electron acceptor for nitrification (e.g. Dunn et al., 2012). Nitrifiers are known to be outcompeted for oxygen by heterotrophic and other chemoautotrophic bacteria under oxygen limiting conditions (Caffrey et al., 1993; Sloth et al., 1995). Therefore, the overall effect of oxygen limitation is that the NH+4 regenerated during organic matter mineralisation in the

Please cite this article as: D. Nizzoli, D. T. Welsh and P. Viaroli, Denitrification and benthic metabolism in lowland pit lakes: The role of trophic conditions, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134804

12

D. Nizzoli et al. / Science of the Total Environment xxx (xxxx) xxx

Fig. 6. Theoretical Nr assimilation and denitrification rates in the littoral zone and water column of the five lakes. Error bars for benthic rates indicate the standard error (n = 3).

sediment, rather than being oxidised and subsequently eliminated via denitrification, effluxes from the sediment to the water column, contributing to the internal nitrogen load. 4.3. Autotrophy and heterotrophy in littoral sediments and the effect of microphytobenthos on benthic nitrogen cycle Chl-a concentrations in the surface sediments and an average net community production of 1.3 ± 0.7 mmol m2 h1 suggest an active microalgal community colonized the littoral sediments of the oligo (CM, CS) and mesotrophic (LV) lakes, which were highly to net autotrophic. Whereas, in the eutrophic lakes (IG, BL) gross productivity was lower and the benthos was net heterotrophic, and an oxygen sink under both light and dark conditions. The shifts in benthic trophic state category reflect a shift in dominance of autotrophic and heterotrophic processes (Rizzo et al., 1996), which largely resulted from a decrease in the photosynthetic activity of benthic microalgae with trophic state. The development of phytoplankton blooms in the water column can reduce the quantity and quality of light reaching the benthos and limit benthic primary production (Vadeboncoeur et al., 2003). In addition, under the more eutrophic conditions Nr assimilation by phytoplankton reduces water column NO 3 concentrations (Table 2) and its availability for benthic communities. Pelagic primary production estimated from the measured water column Chl-a data (see Section 2.6) varied from 9 to 96 mmol C m2 d1 across the five lakes, which corresponded to nitrogen demands to support this production of 2–15 mmol N m2 d1 with the highest values estimated for the more eutrophic lakes and the lowest for the oligotrophic lakes (Fig. 6). The presence of benthic microalgae and primary production in shallow aquatic sediments complicates benthic nitrogen cycling (Dunn et al., 2012; Nizzoli et al., 2014; Sundbäck et al., 2000). During daylight hours, photoassimilation of NH+4 by the microalgae competes with nitrification for NH+4, limiting nitrification and therefore DN (Risgaard-Petersen, 2003). Whereas, under nitrogen replete conditions photosynthetic oxygen production can stimulate DN by enhancing nitrification rates (An and Joye, 2001; Dunn et al., 2012). Conversely, photosynthetic oxygen evolution in the surficial sediment increases oxygen penetration into the sediment limiting DW by increasing the diffusional path length of NO 3 to the denitrification zone (Revsbech et al., 2005). In this study, the dominant influence of benthic microalgae was through oxygen production affecting the sediment oxygen penetration depth and the

diffusion rate of NO 3 to the denitrification zone, as rates of DW were significantly higher under dark compared to light conditions at the littoral sampling sites. Nr assimilation by microalgae also represents a temporal storage of nitrogen in the benthos and modify the source/sink dynamics of the sediment as DIN from the pore or overlying water is converted into organic nitrogen sequestered in the algal biomass (McGlathery et al., 2007; Nizzoli et al., 2014; Sundbäck et al., 2000). Net daily Nr assimilation rates estimated from net benthic primary production (see Section 2.6) were between 0 and 7 mmol N m2 d1 (average 2.25 ± 2.60 mmol N m2 d1) and were comparable to denitrification rates in littoral sediments. The high Nr demand for benthic microalgal production played an important role in regulating NH+4 fluxes across the sediment-water interface. During light incubations, NH+4 effluxes were on average 70% lower and significantly different to those during dark incubations. However, due to the high water column NO 3 concentrations in all lakes, overall sediment-water column DIN fluxes were mainly on average dependent upon NO 3 uptake by the sediment and denitrification. + Yet, the relative contributions of NO 3 and NH4 to overall DIN fluxes varied across the study lakes with their changing trophic status + and BTSI, as sediment NO 3 uptake declined and NH4 efflux increased with increasing eutrophication and BTSI. Denitrification and Nr assimilation by benthic microalgae were the major benthic processes that consumed DIN (Fig. 2S Supplementary material) with the combined amount of Nr consumed decreasing with BTSI from ~10 to ~3 mmol m2 d1 from net autotrophic to net heterotrophic littoral sites (Fig. 6). At autotrophic sites on average only ~55% of the total Nr consumed by denitrification and estimated Nr assimilation was sustained by DIN subtracted from the water column indicating that both denitrification (DN) and benthic microalgae were partially dependent on Nr mineralized within the sediment. Thus, both processes act as eutrophication buffers which limit the recycling of bioavailable nitrogen to the overlying water column by either eliminating it as gaseous end-products or sequestering it as microalgal biomass. The relative contributions of denitrification and Nr assimilation to this buffering capacity varied with the trophic status and BTSI of the individual lakes, with denitrification accounting for from 100% of combined denitrification and Nr assimilation in the more eutrophic, net heterotrophic lakes to ~40% of the total in the oligotrophic, net autotrophic lakes (Fig. 6). These values on the relative role of denitrification are in agreement with the contributions calculated in other freshwater aquatic systems where denitrification accounted for 40–80% of Nr removal due to denitrification and Nr assimilation (Matheson et al., 2002; Kreiling et al., 2011; Nizzoli et al., 2014). However, these results contrast starkly with those observed in autotrophic shallow coastal marine aquatic environments colonized by benthic microalgae, where rates of denitritrification are typically much lower than Nr assimilation because of competition between the benthic primary producers, and nitrifying and denitrifying bacteria for DIN (Dunn et al., 2012; Risgaard-Petersen, 2003; Sundbäck et al., 2000; Welsh et al., 2000). This difference is likely related to the different sources of NO 3 fuelling denitrification in the two environments. Denitrification is mostly dependent on water column NO 3 in the investigated lakes as a consequence of higher water column NO 3 concentrations, whereas in coastal systems denitrification is primarily dependent upon benthic nitrification because of lower water column NO 3 (Sundbäck et al., 2000; Dunn et al., 2012). In summary, our results expand upon previous measured rates in littoral areas of pit lakes (Nizzoli et al., 2014). In shallow aquatic ecosystems where light and DIN availability are not limiting, benthic denitrification and Nr assimilation play important and comparable roles in nitrogen removal. However, with increasing trophic status the magnitude of these Nr processing rates declines, reduc-

Please cite this article as: D. Nizzoli, D. T. Welsh and P. Viaroli, Denitrification and benthic metabolism in lowland pit lakes: The role of trophic conditions, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134804

D. Nizzoli et al. / Science of the Total Environment xxx (xxxx) xxx

ing the total amount of nitrogen consumed by the sediment compartment, and primary production and Nr assimilation in pelagic habitats becomes more important (Fig. 6). In particular denitrification declines with increasing trophic status likely because under the more eutrophic conditions Nr assimilation by phytoplankton reduces water column NO 3 concentrations, and its availability for denitrifying bacteria in the sediment. Additionally, increased phytoplankton production in the water column, reduces light availability at the sediment surface, limiting benthic primary production and associated Nr assimilation into primary producer biomass. This results in an increasing proportion of the Nr produced during organic matter mineralisation in the sediment effluxing to the water column. 4.4. Management and design of pit lakes to improve nitrogen removal In this study, nitrogen removal rates due to denitrification and Nr assimilation in the five studied pit lakes, especially in the littoral zone, are among the highest reported in the literature, with measured denitrification rates corresponding to net Nr removal of 0.3–39 (hypolimnion) and 36–98 kg N km2 d1 (littoral zone) which are comparable to the estimated Nr surplus (between 2 and 85 t N km2 y1) from agricultural practises in the watershed (Viaroli et al., 2018). Therefore, despite their typically small size, these lakes can play a role in mitigating Nr loads from agriculture and act as nutrient filters to reduce Nr loads to downstream ecosystems. Differences in properties of the shallow littoral areas and deep hypolimnetic sediments of the lakes with respect to Nr removal mean that pit lakes could be designed by shaping the features of the waterbody to maximise their capacity to mitigate nutrient loads. For example, if we assume that the measured denitrification rates are representative of the littoral and hypolimnetic zones of each lake, then total annual Nr removal by denitrification would be between 0.4 and 2.2 t N y1 depending on the size and trophic state of the lake. However, between 30 and 95% of total nitrogen removal per lake would be accounted for by the littoral sediment zone, despite this zone representing only 9 to 24% of the total surface area of the individual lakes (Table 1). Typically, however, pit lakes, such as those investigated in this study, have a high depth-to-surface area ratio with relatively steep sloping sides, as this maximises the cost to benefit ratio of sand/gravel quarrying (Blanchette and Lund, 2016; Søndergaard et al., 2018). Consequently, shallow littoral areas are limited to a narrow band around the edge of the pit lake and represent only a minor fraction of the total lake area. Our results show that this business as usual approach diminishes the potential nutrient filter function of the pit lake, as littoral areas represent stronger Nr sinks than the deep sediment zones. Thus, the nutrient filtering capacity of these artificial lakes could be enhanced at the design stage, after the cessation of quarrying, by maximising the relative area of the littoral zone, for example by excavating shallow areas around all or part of the quarry before it is flooded. This would be a desirable outcome for both better floodplain Nr management and in slowing or preventing eutrophication of the resultant lake by increasing the degree of Nr sequestration in benthic primary producer biomass, and especially Nr loss via denitrification. Additionally, increased extension of littoral areas would favour colonisation by submerged macrophytes that would contribute to Nr management by reducing N-mobility and enhancing N-removal via denitrification (Nizzoli et al., 2014). 5. Conclusions In summary, our results highlight that floodplain pit lakes, can provide the ecosystem services formerly supplied by natural wet-

13

lands, i.e. nitrogen uptake and removal. This function could be improved by designing and managing these lakes to maximise their N filtering capacity. Benthic Nr retention rates are higher in the littoral zone and decrease as the lake trophic state increases. Overall, under eutrophic conditions, the capacity of the benthic system to provide Nr retention and loss is reduced and the benthic compartment shifts from being a net N-sink in oligo mesotrophic lakes to a net source of DIN to the water column in the most eutrophic lakes. Consequently, increased dominance of pelagic production and Nr assimilation with trophic state acts as a feedback loop that accelerates lake eutrophication. Thus, care needs to be taken in the design and management of pit lakes to increase the extension of shallow littoral areas and to avoid the hypereutrophic and anoxic conditions that could favour nitrogen recycling over dissipation if their nutrient filter role is to be maintained. The latter could favour nitrogen recycling over dissipation, deteriorating the capacity of nutrient filtering and, ultimately, the related regulatory service. 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. Acknowledgements This study was funded by Lombardy Region within the project ‘‘Assessment of the limnological status of the Lombardy Region lakes and of nutrients loads formation in heavily exploited watersheds” and by Bassanetti s.r.l. (Monticelli d’Ongina, Piacenza). The authors are very grateful to Daniele Magni and Marco Parini (Lombardy Region) for their support during the research activities, to Emanuele Emani and Claudio Bassanetti who supported the organisation of this study, and to Daniele Longhi that supported the field activities.

Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi.org/10.1016/j.scitotenv.2019.134804. References An, S., Joye, S.B., 2001. Enhancement of coupled nitrification-denitrification by benthic photosynthesis in shallow estuarine sediments. Limnol. Oceanogr. 46, 62–74. APHA (American Public Health Association), 1998. Standard Methods for the Examination of Water and Wastewaters. APHA, Washington, DC. Aspila, K.I., Agemian, H., Chau, A.S.Y., 1976. A semi-automated method for the determination of inorganic, organic and total phosphate in sediments. Analyst 101 (1200), 187–197. Azzoni, R., Nizzoli, D., Bartoli, M., Christian, R.R., Viaroli, P., 2015. Factors controlling benthic biogeochemistry in urbanized coastal systems: an example from Venice (Italy). Estuaries Coasts 38 (3), 1016–1031. Billen, G., Garnier, J., Lassaletta, L., 2013. The nitrogen cascade from agricultural soils to the sea: modelling nitrogen transfers at regional watershed and global scales. Philos. Trans. R. Soc. Lond. B 368, 20130123. Blanchette, M.L., Lund, M.A., 2016. Pit lakes are a global legacy of mining: an integrated approach to achieving sustainable ecosystems and value for communities. Curr. Opin. Environ. Sustain. 23, 28–34. Bruesewitz, D.A., Hamilton, D.P., Schipper, L.A., 2011. Denitrification potential in lake sediment increases across a gradient of catchment agriculture. Ecosystems 14 (3), 341–352. Bruesewitz, D.A., Tank, J.L., Hamilton, S.K., 2012. Incorporating spatial variation of nitrification and denitrification rates into whole-lake nitrogen dynamics. Journal of Geophysical Research. Biogeosciences 117, G00N07. 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 (2), 89–96.

Please cite this article as: D. Nizzoli, D. T. Welsh and P. Viaroli, Denitrification and benthic metabolism in lowland pit lakes: The role of trophic conditions, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134804

14

D. Nizzoli et al. / Science of the Total Environment xxx (xxxx) xxx

Caffrey, J.M., Sloth, N.P., Kaspar, H.F., Blackburn, T.H., 1993. Effect of organic matter loading on nitrification and denitrification in a marine sediment mesocosm. FEMS Microbiol. Ecol. 12, 159–167. Canfield, D. E., Kristensen, E., Thamdrup, B., 2005. The nitrogen cycle. In: Advances in Marine Biology. Aquatic Geomicrobiology. Academic Press, 2005. p. 205-267463. Castaldelli, G., Soana, E., Racchetti, E., Vincenzi, F., Fano, E.A., Bartoli, M., 2015. Vegetated canals mitigate nitrogen surplus in agricultural watersheds. Agric. Ecosyst. Environ. 212, 253–262. Cheng, F.Y., Basu, N.B., 2017. Biogeochemical hotspots: Role of small water bodies in landscape nutrient processing. Water Resour. Res. 53 (6), 5038–5056. Dalsgaard, T., Nielsen, L.P., Brotas, V., Viaroli, P., Underwood, G.J.C., Nedwell, D.B., Sundbäck, K., Rysgaard, S., Miles, A., Bartoli, M., Dong, L., Thornton, D.C.O., Ottosen, L.D.M., Castaldelli, G., Risgaard-Petersen, N., 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, Silkeborg, Denmark, p. 62. den Heyer, C., Kalff, J., 1998. Organic matter mineralization rates in sediments: a within- and among-lake study. Limnol. Oceanogr. 43 (4), 695–705. Dunn, R.J.K., Welsh, D.T., Jordan, M.A., Waltham, N.J., Lemckert, C.J., Teasdale, P.R., 2012. Benthic metabolism and nitrogen dynamics in a sub-tropical coastal lagoon: microphytobenthos stimulate nitrification and nitrate reduction through photosynthetic oxygen evolution. Estuar. Coast. Shelf Sci. 113, 272– 282. Finlay, J.C., Small, G.E., Sterner, R.W., 2013. Human influences on nitrogen removal in lakes. Science 342 (6155), 247–250. Galloway, J.N., Aber, J.D., Erisman, J.W., Seitzinger, S.P., Howarth, R.W., Cowling, E.B., Cosby, B.J., 2003. The nitrogen cascade. Bioscience 53, 341–356. Gardner, W.S., McCarthy, M.J., 2009. Nitrogen dynamics at the sediment–water interface in shallow, sub-tropical Florida Bay: why denitrification efficiency may decrease with increased eutrophication. Biogeochemistry 95 (2–3), 185– 198. Glibert, P.M., 2017. Eutrophication, harmful algae and biodiversity – challenging paradigms in a world of complex nutrient changes. Mar. Pollut. Bull. 124 (2), 591–606. 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. Harrison, J.A., Maranger, R.J., Alexander, R.B., Giblin, A.E., Jacinthe, P.A., Mayorga, E., Seitzinger, S.P., Sobota, D.J., Wollheim, W.M., 2009. The regional and global significance of nitrogen removal in lakes and reservoirs. Biogeochemistry 93 (1– 2), 143–157. Helmer, C., Labroue, L., 1993. Denitrification in gravel-pit lakes. Hydrobiologia 252 (1), 35–44. Herrman, K.S., White, J.R., 2008. Denitrification in intact sediment cores from a constructed wetland: examining the isotope pairing technique. Appl. Geochem. 23 (8), 2105–2112. Howarth, R.W., Marino, R., 2006. Nitrogen as the limiting nutrient for eutrophication in coastal marine ecosystems: evolving views over three decades. Limnol. Oceanogr. 51 (1part2), 364–376. Kattner, E., Schwarz, D., Maier, G., 2000. Eutrophication of gravel pit lakes which are situated in close vicinity to the river Donau: water and nutrient transport. Limnol. Ecol. Manage. Inland Waters 30 (3), 261–270. Koroleff, F., 1970. Direct Determination of Ammonia in Natural Waters as Indophenol Blue. Information on Techniques and Methods for Seawater Analysis. I.C.E.S Interlaboratory Rep. No. 3, pp. 19–22. Kreiling, R.M., Richardson, W.B., Cavanaugh, J.C., Bartsch, L.A., 2011. Summer nitrate uptake and denitrification in an upper Mississippi River backwater lake: the role of rooted aquatic vegetation. Biogeochemistry 104 (1–3), 309–324. Lenth, R., Singmann, H., Love, J., Buerkner, P., Herve, M., 2018. Estimated Marginal Means, Aka Least-Squares Means v 1.3.1. Lorenzen, C.J., 1967. Determination of chlorophyll and pheopigments: spectrophotometric equations. Limnol. Oceanogr. 12 (2), 343–346. Matheson, F.E., Nguyen, M.L., Cooper, A.B., Burt, T.P., Bull, D.C., 2002. Fate of 15Nnitrate in unplanted, planted and harvested riparian wetland soil microcosms. Ecol. Eng. 19 (4), 249–264. McCrackin, M.L., Elser, J.J., 2010. Atmospheric nitrogen deposition influences denitrification and nitrous oxide production in lakes. Ecology 91 (2), 528– 539. McGlathery, K.J., Sundbäck, K., Anderson, I.C., 2007. Eutrophication in shallow coastal bays and lagoons: the role of plants in the coastal filter. Mar. Ecol. Prog. Ser. 348, 1–18. Mollema, P.N., Antonellini, M., 2016. Water and (bio)chemical cycling in gravel pit lakes: a review and outlook. Earth Sci. Rev. 159, 247–270. Muellegger, C., Weilhartner, A., Battin, T.J., Hofmann, T., 2013. Positive and negative impacts of five Austrian gravel pit lakes on groundwater quality. Sci. Total Environ. 443, 14–23. National Institute of Statistic, 2017. Le attività estrattive da cave e miniere (Extraction activities from pits and mines) (In italian). Nielsen, L.P., 1992. Denitrification in sediment determined from nitrogen isotope pairing. FEMS Microbiol. Lett. 86 (4), 357–362. Nizzoli, D., Carraro, E., Nigro, V., Viaroli, P., 2010. Effect of organic enrichment and thermal regime on denitrification and dissimilatory nitrate reduction to ammonium (DNRA) in hypolimnetic sediments of two lowland lakes. Water Res. 44 (9), 2715–2724.

Nizzoli, D., Welsh, D.T., Longhi, D., Viaroli, P., 2014. Influence of Potamogeton pectinatus and microphytobenthos on benthic metabolism, nutrient fluxes and denitrification in a freshwater littoral sediment in an agricultural landscape: N assimilation versus N removal. Hydrobiologia 737 (1), 183–200. Nizzoli, D., Bartoli, M., Azzoni, R., Longhi, D., Castaldelli, G., Viaroli, P., 2018. Denitrification in a meromictic lake and its relevance to nitrogen flows within a moderately impacted forested catchment. Biogeochemistry 137 (1–2), 143– 161. Peckenham, J.M., Thornton, T., Whalen, B., 2009. Sand and gravel mining: effects on ground water resources in Hancock county, Maine, USA. Environ. Geol. 56 (6), 1103–1114. Pina-Ochoa, E., Álvarez-Cobelas, M., 2006. Denitrification in aquatic environments: a cross-system analysis. Biogeochemistry 81 (1), 111–130. Pinheiro, J., Bates, D., DebRoy, S., Sarkar, D., R Core Team. 2016. nlme: Linear and Nonlinear Mixed Effects Models. R Package Version 3.1-126. R Core Team, 2018. R: A Language and Environment for Statistical Computing. R Foundation for Statistical Computing, Vienna, Austria. http://www.R-project. org/. Revsbech, N., Jacobsen, J., Nielsen, L., 2005. Nitrogen transformations in microenvironments of river beds and riparian zones. Ecol. Eng. 24 (5), 447–455. Risgaard-Petersen, N., Rysgaard, S., 1995. Nitrate reduction in sediments and waterlogged soil measured by 15N techniques. In: Methods in applied soil microbiology. Academic Press, London, pp. 1–13. Risgaard-Petersen, N., 2003. Coupled nitrification-denitrification in autotrophic and heterotrophic estuarine sediments: on the influence of benthic microalgae. Limnol. Oceanogr. 48 (1), 93–105. Risgaard-Petersen, N., Nielsen, L.P., Dalsgaard, T., Meyer, R.L., 2003. Application of the isotope pairing technique in sediments where anammox and denitrification coexist. Limnol. Oceanogr. Methods 1, 63–73. Rissanen, A.J., Tiirola, M., Hietanen, S., Ojala, A., 2013. Interlake variation and environmental controls of denitrification across different geographical scales. Aquat. Microb. Ecol. 69 (1), 1–16. Rizzo, W.M., Dailey, S.K., Lackey, G.J., Christian, R.R., Berry, B.E., Wetzel, R.L., 1996. A metabolism-based trophic index for comparing the ecological values of shallow-water sediment habitats. Estuaries 19 (2), 247–256. Rysgaard, S., Risgaard-Petersen, N., Sloth, N.P., Kim, J., Nielsen, L.P., 1994. Oxygen regulation of nitrification and denitrification in sediments. Limnol. Oceanogr. 39 (7), 1643–1652. Saunders, D.L., Kalff, J., 2001. Denitrification rates in the sediments of Lake Memphremagog, Canada–USA. Water Res. 35 (8), 1897–1904. Schramm, W., Gualberto, E., Orosco, C., 1984. Release of dissolved organic matter from marine tropical reef plants: temperature and desiccation effects. Bot. Mar. 27, 71–77. Schullehner, J., Hansen, B., Thygesen, M., Pedersen, C.B., Sigsgaard, T., 2018. Nitrate in drinking water and colorectal cancer risk: a nationwide population-based cohort study. Int. J. Cancer 143 (1), 73–79. Scott, J.T., McCarthy, M.J., Gardner, W.S., Doyle, R.D., 2008. Denitrification, dissimilatory nitrate reduction to ammonium, and nitrogen fixation along a nitrate concentration gradient in a created freshwater wetland. Biogeochemistry 87 (1), 99–111. Seitzinger, S., Harrison, J.A., Böhlke, J.K., Bouwman, A.F., Lowrance, R., Peterson, B., Tobias, C., Van Drecht, G., 2006. Denitrification across landscapes and waterscapes: a synthesis. Ecol. Appl. 16 (6), 2064–2090. Sloth, N.P., Blackburn, T.H., Hansen, L.S., Risgaard-Petersen, N., Lomstein, B.A., 1995. Nitrogen cycling in sediments with different organic loading. Mar. Ecol. Prog. Ser. 116, 63–70. Small, G.E., Cotner, J.B., Finlay, J.C., Stark, R.A., Sterner, R.W., 2014. Nitrogen transformations at the sediment–water interface across redox gradients in the Laurentian Great Lakes. Hydrobiologia 731 (1), 95–108. Søndergaard, M., Lauridsen, T.L., Johansson, L.S., Jeppesen, E., 2018. Gravel pit lakes in Denmark: chemical and biological state. Sci. Total Environ. 612, 9–17. Sutton, M.A., Howard, C.M., Erisman, J.W., Billen, G., Bleeker, A., Grennfelt, P., van Grinsven, H., Grizzetti, B. (Eds.), 2011. The European Nitrogen Assessment. Cambridge University Press, Cambridge, pp. 271–297. Sundbäck, K., Miles, A., Göransson, E., 2000. Nitrogen fluxes, denitrification and the role of microphytobenthos in microtidal shallow-water sediments: an annual study. Mar. Ecol. Prog. Ser. 200, 59–76. Swaney, D.P., Hong, B., Ti, C., Howarth, R.W., Humborg, C., 2012. Net anthropogenic nitrogen inputs to watersheds and riverine N export to coastal waters: a brief overview. Curr. Opin. Environ. Sustain. 4 (2), 203–211. Tavernini, S., Nizzoli, D., Rossetti, G., Viaroli, P., 2009. Trophic state and seasonal dynamics of phytoplankton communities in two sand-pit lakes at different successional stages. J. Limnol. 68 (2), 217–228. Townsend, A.R., Howarth, R.W., Bazzaz, F.A., Booth, M.S., Cleveland, C.C., Collinge, S. K., Dobson, A.P., Epstein, P.R., Keeney, D.R., Mallin, M.A., Rogers, C.A., Wayne, P., Wolfe, A.H., 2003. Human health effects of a changing global nitrogen cycle. Front. Ecol. Environ. 1 (5), 240–246. UEPG, 2017. European Aggregates Association. Annual Review 2015-2016e. http:// www.uepg.eu/uploads/Modules/Publications/uepg-ar2016-17_32pages_v04_ small.pdf. Uusheimo, S., Tulonen, T., Aalto, S.L., Arvola, L., 2018. Mitigating agricultural nitrogen load with constructed ponds in northern latitudes: a field study on sediment denitrification rates. Agric. Ecosyst. Environ. 261, 71–79. Vadeboncoeur, Y., Jeppesen, E., Zanden, M., Schierup, H.H., Christoffersen, K., Lodge, D.M., 2003. From Greenland to green lakes: cultural eutrophication and the loss of benthic pathways in lakes. Limnol. Oceanogr. 48 (4), 1408–1418.

Please cite this article as: D. Nizzoli, D. T. Welsh and P. Viaroli, Denitrification and benthic metabolism in lowland pit lakes: The role of trophic conditions, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134804

D. Nizzoli et al. / Science of the Total Environment xxx (xxxx) xxx Valderrama, J.C., 1981. The simultaneous analysis of total nitrogen and total phosphorus in natural waters. Mar. Chem. 10, 109–122. Verstraete, W., Focht, D.D., 1977. Biochemical ecology of nitrification and denitrification. In Advances in Microbial ecology Springer, Boston, MA, pp. 135–214. Viaroli, P., Soana, E., Pecora, S., Laini, A., Naldi, M., Fano, E.A., Nizzoli, D., 2018. Space and time variations of watershed N and P budgets and their relationships with reactive N and P loadings in a heavily impacted river basin (Po river, Northern Italy). Sci. Total Environ. 639, 1574–1587. Vollenweider, R.A., Kerekes, J., 1982. Eutrophication of Waters. Monitoring Assessment and Control. Organization for Economic Co-Operation and Development (OECD), Paris. Weilhartner, A., Muellegger, C., Kainz, M., Mathieu, F., Hofmann, T., Battin, T.J., 2012. Gravel pit lake ecosystems reduce nitrate and phosphate concentrations in the outflowing groundwater. Sci. Total Environ. 420, 222–228.

15

Welsh, D.T., Bartoli, M., Nizzoli, D., Castaldelli, G., Riou, S.A., Viaroli, P., 2000. Denitrification, nitrogen fixation, community primary productivity and inorganic-N and oxygen fluxes in an intertidal Zostera noltii meadow. Mar. Ecol. Prog. Ser. 208, 65–77. Welsh, D.T., Castadelli, G., Bartoli, B., Poli, D., de Wit, R., Careri, M., Viaroli, P., 2001. Denitrification in an intertidal seagrass meadow, a comparison of 15N-isotope and acetylene block techniques: dissimilatory nitrate reduction to ammonium as a source of N2O? Mar. Biol. 139, 1029–1036. Wetzel, R.G., Likens, G.E., 1991. Limnological Analyses. Springer-Verlag, New York, p. 391. Wetzel, R.G., 2001. Limnology. Lake and River Ecosystems. Academic Press, London, p. 1006. White, P.A., Kalff, J., Rasmussen, J.B., Gasol, J.M., 1991. The effect of temperature and algal biomass on bacterial production and specific growth rate in freshwater and marine habitats. Microb. Ecol. 21, 99–118.

Please cite this article as: D. Nizzoli, D. T. Welsh and P. Viaroli, Denitrification and benthic metabolism in lowland pit lakes: The role of trophic conditions, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134804