Study site

The Teagasc Johnstown Castle, Co. Wexford, (52°17'30″N, 6°29'50″W) (Fig 1) study site comprises two units of grassland, one of 72.9 ha, termed “down-gradient”, and an “up-gradient” of 50.8 ha separated by a roadway. The up-gradient subsurface drainage system is connected to the down-gradient system through an underground connector pipe. The farm, considering the two grassland units together, is intensive and operates at 3.1 Livestock units per hectare (LU ha^-1). Nitrogen inputs arise from urea (spread February to April), calcium ammonium nitrate (CAN) (May to September) and manure (spring) for a total of 257 kg N ha^-1 inputs of inorganic fertiliser and 103 kg N ha^-1 of organic (5 years average). The central area of the dairy farm receives dairy soiled water (DSW, consisting of rainwater, yard and milk parlour washings) from Feb to Oct through a rotational irrigation system (Fig 1, former spreading area: plot with locations 11 and 14; current spreading area: between the Met station and location 19). To quantify the N surplus that could affect the drainage system, farm partial N balances for each year were calculated as per Treacy et al. [17], utilising stocking rates, N inputs (inorganic and organic fertilisers and concentrate feed (volume and composition)) and N exports (milk production (volume and composition) and slurry). Cows were milked twice daily (07·30 h and 15·30 h), with milk yield (kg) registered for each cow each time. The milk composition (fat, protein and lactose concentrations) for each cow was tested on a successive morning and evening every two weeks using a Milkoscan 203 (Foss Electric DK-3400, Hillerød, Denmark). Milk solids were calculated using the method of Tyrell and Reid [18]. For both milk and concentrates fed the N value is averaged across the farm. This region has an average annual air temperature of 10°C (1981–2011), average annual rainfall (1981–2011) is 1037.5 mm with maximum intensity between Sep and Nov. There is an in situ synoptic meteorological station (Fig 1) on site which records daily rainfall, wind speed and hours of sunshine. This data was inputted into the grassland hybrid soil moisture deficit (HSMD) model of Schulte et al. [19] to estimate effective drainage (ED, mm day^-1). To examine differences in ED modelling was conducted for well-, moderately- and poorly-drained soils using data from 2008 to 2017. Soil texture varies from fine loam to clay loam (Brown Earth, Gleyic Cambisol with Irish Sea till origin [20], Fig 1), with small areas of sandy textured soils. Subsoil is of moderate permeability (0.2–10 m; 5 x 10⁻⁸ m s^-1 <k_s< 5 x 10⁻⁴ m s^-1) but can be interspersed with high permeability gravel and/or sand lenses [8, 21]. Unsaturated zone hydrologic time lag is from 1 (shallowest areas) to 3 (deepest areas) years with probability analysis indicating 1.5 years 85% of the time [14], presenting a mismatch between best management practice at farm level, when nutrients are lost, stored and mineralised and when leaching affects N concentrations along subsurface pathways. The average water table depth on site from 2005–2014 is 2.8 (±1.7) m below ground level (bgl) (deepest from July to Sep i.e. 3.0 m bgl and shallowest from December to February i.e. 2.3 m bgl). The low permeability bedrock is Pre-Cambrian greywacke mixed with quartzite at a depth of 10–12 m (k_s, 3.6 x 10⁻⁶ m s^-1), containing a poorly productive aquifer, which is classified as receiving 1 to 50 mm yr^-1 [14].

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Fig 1. Map of the intensive farm merging soil texture, drainage class, position of the surface and subsurface drainage networks including the lake system and outlet and groundwater monitoring locations (Squares indicate piezometers, multilevel boreholes are indicated by triangles, end-of-pipe locations by circles and open ditch locations by asterisks).

https://doi.org/10.1371/journal.pone.0219479.g001

Surface and subsurface monitoring network

A large surface and subsurface monitoring network was collated in GIS and a map developed for the purpose of the study. The open ditch network and groundwater monitoring components of the system were well documented, but the subsurface drainage system remained unmapped. Examination of historical maps and discussions with farm staff identified likely positions of in-field drains, which were confirmed by field work (Fig 1, S1 Table). Both up- and down-gradient units contain poorly-imperfectly drained soils (79% and 28%, respectively) (Fig 1). These areas are all artificially drained by an in-field pipe network of 10.1 km (Fig 1) composed of either corrugated slotted plastic (80 or 100 mm at 0.5–1 m depth) or concrete non-perforated pipes (600 mm at 1 m depth act as connectors).

The components of the surface and subsurface drainage system are as follows: 1) Up-gradient component: a herring bone drainage network (80 mm, variable depth) conveys drainage water from an area of 24 ha, to an underground outlet (D2, Fig 1) and joins with another outlet at D3, which passes to a junction also fed by a drainage system passing through D1. This composite migrates in a fully cased concrete pipe and eventually discharges into the open ditch in the down-gradient unit. 2) Down-gradient component including open ditch: the subsurface drain that transfers water from locations 2, 3, 7, 8, 9 and 10 is an ad-hoc drain developed over time with many unknown extensions and discharges to the start of the open ditch. A short herring bone system discharges water into the open ditch adjacent to location 14. The open ditch extends from 11 to D4, before being piped underground to the outlet at D8 (Fig 1). There is an access point in the underground section at D5-6-7. Individual drains connect to the underground part of this primary drain at location 30 (fully cased draining a marl pond) and 36. 3) Unconnected components: an in-field herring bone drain in the area of 15, 16 and 17 flows in the opposite direction to the open drain and has an offsite outlet, as does the system around 19 and 20. Any discharge from the nearby artificial lake system (D9) does not affect D8. Prior to the 2014 sampling campaign, water sampling locations were added to this system, i.e. three positions along the subsurface drainage system (D1, D2, D6), one position along the surface stream (D5) and three within the groundwater network (1, 22, 23) (Fig 1).

Historical data

Twelve years of data for the down gradient dairy farm were collated from the groundwater monitoring network (S1 Table, No 5). The groundwater monitoring system consists of three components: 1) five sets of multilevel boreholes (Fig 1) representing three depths: subsoil (4.0–7.5 m bgl), bedrock (16.8–23.0 m bgl) and the subsoil-bedrock interface (11.0–13.0 m bgl); 2) a network of piezometers and boreholes (Fig 1), with screen depths from 1.95 to 8.95 m bgl installed to sample shallow groundwater within the subsoil layer; 3) a single borehole drilled to 37 m bgl was installed to sample deep groundwater (18, Fig 1).

Fieldwork was conducted from Dec 2005 to June 2017 (S1 Table, No 5). During this period, grab samples were taken from open ditch and end-of-pipe locations indicated in Fig 1. Groundwater samples were collected after purging three volumes [22] with a peristaltic pump (Model 410, Solinst Canada Ltd.) and teflon outlet, connected to a flow cell fitted with an In-Situ Multi-parameter Probe (pH, temperature (T), electrical conductivity (EC), DO and Eh) (In Situ Inc., USA). Duplicate 50 ml samples were collected in HDPE screw top bottles and filtered with 0.45 μm cellulose acetate filters (Sartorius Stedim Biotech GmbH, Germany). The water table depth was measured with an electronic dipper (Van Walt Ltd., Surrey, UK).

Current data

Fieldwork was undertaken in Sep 2014 and June 2017 to collect water samples for all locations in Fig 1. These two sampling campaigns were carried out to integrate different techniques (i.e. stable isotope and dissolved gas analyses (only for Sep 2014) and all elements of the water continuum (i.e. shallow and groundwater, in field drainage and open ditches). These campaigns also deduced the stability of the N signatures and dissolved gases on site within shallow water which was highlighted by Baily et al. [14] and Jahangir et al. [10,11] across the multilevel borehole network (S1 Table). Water was sampled using a bladder pump (flow rate of 100 ml min^-1) with a Teflon outlet tube (diameter: 0.6 cm) (Geotech Environmental Equipment, Inc., USA) following a low-flow micro-purging protocol [22]. This pump minimises sample mixing and degassing [10]. Otherwise a peristaltic pump and 20 ml syringe connected to a Teflon tube (diameter: 0.5 cm) with 3-way stopcock was used. Triplicate 50 ml surface water samples from pipes or open drains were collected manually in HDPE screw top bottles and stored at 4°C until analysed. One replicate was filtered in the field (0.45 μm cellulose acetate filter). As per historical data an electronic dipper and In-Situ Multi-parameter Probe were used in the field.

Water and gas analyses

Water samples (both current and historical) were analysed within two weeks of collection for the following: NO₂^--N, NH₄⁺-N, Total Oxidised Nitrogen (TON) and dissolved organic carbon (DOC) were quantified using an Aquakem 600 Discrete Analyser (Aquakem 600A, 01621 Vantaa, Finland). Concentrations of NO₃^--N were calculated by subtraction of NO₂^--N from TON (NO₃^--N + NO₂^--N). Total Nitrogen (TN) was determined by alkaline persulfate oxidation [23].

For current fieldwork, duplicate water samples were collected (all locations) for excess-N₂ estimation in 12 ml exetainers (LabcoWycomb Ltd., UK) after overflow of 10 ml and duplicate groundwater samples were collected for the quantification of dissolved N₂O, carbon dioxide (CO₂) and methane (CH₄). The analyses and calculation of gaseous concentrations for excess-N₂, dissolved N₂O, CO₂ and CH₄ were carried out as per Clagnan et al. [5,6]. The indirect N₂O-N emission factor (N₂O-N EF₅g) for groundwater was calculated from dissolved-N₂O and N-inputs, as per Weymann et al. [24] using the following equation: EF₅g(1) = (N₂O-N)/(N₂O-N + N₂-N + NH₄⁺-N + NO₃^--N + NO₂^--N + DON).

Stable isotope analysis

In Sep 2014 and June 2017 groundwater samples (40 ml) were collected on two dates, filtered in the field through 0.2 μm polyethersulfone filters (Sartorius Stedim Biotech GmbH, Germany) and stored at -20°C in 50 ml polyethylene screw cap tubes. Samples were analysed (Dept. of Catchment Hydrology, UFZ, Germany) for the isotopic composition of NO₃^- (^15/14N and ^18/16O), NH₄⁺ (^15/14N), and H₂O (^2/1H and ^18/16O). Gas exetainers of 12 ml were additionally used for dissolved-N₂O (^15/14N and ^18/16O). Isotope values are reported in δ‰ relative to international standards (AIR for N and VSMOW (Vienna Standard Mean Ocean Water) for O and H). Water δ¹⁸O and δD (δ²H) signatures for H₂O were analysed in accordance with [5,6]. The results are interpreted according to a modified Rayleigh equation: f = 1—e^{(δs–δs0)/ε}, where f is the fraction of substrate remaining, ε is the isotope enrichment factor, and δs and δs0 are the isotopic composition of the residual substrate and initial substrate, respectively. The δ¹⁵N-NH₄⁺ was measured in subsurface locations with detectable NH₄⁺-N concentration. The δ¹⁵N signature for NH₄⁺ was obtained as in Zhang et al. [25]. The δ¹⁵N-NO₃^- and δ¹⁸O-NO₃^- were obtained as in [26]. The δ¹⁵N and δ¹⁸O composition of the produced N₂O (plus that of the dissolved N₂O samples) were measured as in [5].

Statistical analysis

Different methods (t-test, oneway ANOVA and Tukey's HSD test (IBM SPSS Statistics version 24)) were used to determine possible correlations and differences between nutrient, isotopic, and gaseous data and other hydrochemical parameters within the management groups.

Partial N balance

Annual total N-inputs were estimated 364 kg N ha^-1 (296 kg N ha^-1—fertilizers, 68 kg N ha^-1—feed concentrates) whereas national average for intensive farms is 223 kg N ha^-1 ([17]) (Table 1). Irrigation N was composed of slurry and DSW and it was accounted as organic fertilisers. In this area, between 2013–2014, the annual sum of wet and dry deposition was 6 and 8 kg ha^-1 per year (2–4 kg ha^-1 per year of NH₄⁺ for wet and dry deposition, respectively) [27]. Mineralization is a larger N source than ammonia deposition but variable and dependant on soil and climatic conditions. Murphy et al. [28] found that on sandy loam soils, with 6 kg ha^-1 N deposition, net N mineralisation rates of 156 kg ha^-1 per year occurred while of 161 kg ha^-1 per year on clay loam.

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Table 1. Five years annual N balance for the farm.

https://doi.org/10.1371/journal.pone.0219479.t001

With an N-output of 140 kg N ha^-1 exported in milk and slurry, the estimated N-surplus is 224 kg N ha^-1, which is stored in soil, used by the plant, lost from the system through leaching and runoff, or lost in the atmosphere as N₂, N₂O or as gaseous ammonia [29]. Ammonia emissions vary depending on climatic conditions during N application and applied N source; Misselbrook et al. [30] calculated that on English soils the emission factor from slurry and DSW is 34% of total ammonia N applied and 6% of total ammonia N applied, respectively.

Spatial and temporal variation in aqueous N-species

Nitrate varied spatially from 25.3 mg NO₃^--N l^-1 to below detection. The overall average is low e.g. 4.5 mg NO₃^--N l^-1 in 2005, 3.0 mg NO₃^--N l^-1 in 2010, 3.6 mg NO₃^--N l^-1 in 2015 or 3.3 mg NO₃^--N l^-1 in 2017 (S1 Fig). In Sep 2014, location 1 (dominated by low permeability soil) had a NO₃^--N concentration of 3.4 mg NO₃^--N l^-1, whereas in the down-gradient unit (variable soil permeability) three distinct shallow groundwater signatures emerged: a) in the north (locations 4–9), shallow samples ranged from 6.2 to 8.3 mg NO₃^--N l^-1, and deeper layers ranged from 5.7 to 7.0 mg NO₃^--N l^-1; b) well drained soils close to the sampling location 14 (see Fig 1, S2 Table) where concentrations ranged from 4.0 to 6.3 mg NO₃^--N l^-1; and c) central part of the farm (surrounding location 30) on a well-moderately drained soil exhibited concentrations up to 7.6 mg NO₃^--N l^-1 (Fig 2). The deepest well on the farm (location 18, depth: 37 m bgl) had a concentration of 3.07 mg NO₃^--N l^-1 and below NO₃^--N maximum acceptable concentration (MAC), indicating the vertical extent of the NO₃- plume to be around 16 m. There was no elevated concentration of NO₃^--N in boreholes at the south end border of the farm (36–38) on imperfectly drained soil (Figs 1 and 2). At all times NO₃^--N discharging from the drainage system was consistently low at <5.65 mg NO₃^--N l^-1 (the contamination threshold recommended by [31,32]).

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Fig 2. Depth specific N₃^--N concentration as per the sampling campaign in Sep 2014.

Top left: drainage system, top right: 2.95–4.5 m bgl, middle left 4.5–6 m bgl, middle right 6–9 m bgl, bottom left 11–13 m bgl, bottom right: below 16 m bgl.

https://doi.org/10.1371/journal.pone.0219479.g002

Ninety percent of locations had at least one sporadic concentration above the NH₄⁺-N MAC over the years (2005–2017). End of pipe samples from the drainage system averaged at 0.01 mg NH₄⁺-N l^-1. In Sep 2014 the groundwater average was 0.98 mg NH₄⁺-N l^-1 (NH₄⁺-N MAC: 0.23 mg NH₄⁺-N l^-1 [32]) mainly due to three elevated locations. The highest of these 22.7 mg l^-1 occurred at location 37 (S1 Fig). Elevated locations typically occurred > 4 m bgl and at central locations (Fig 3).

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Fig 3. Depth specific NH₄⁺-N concentration as per the sampling campaign in Sep 2014.

Top left: drainage system, top right: 2.95–4.5 m bgl, middle left 4.5–6 m bgl, middle right 6–9 m bgl, bottom left 11–13 m bgl, bottom right: below 16 m bgl.

https://doi.org/10.1371/journal.pone.0219479.g003

Dissolved-N₂O and excess-N₂

Dissolved gas (dissolved-N₂O and excess-N₂) results varied spatially and with depth. In water samples, the excess-N₂ concentration averaged 2.28 mg N₂-N l^-1, with higher values at the interface and bedrock layer, while dissolved N₂O was highest in the subsoil i.e. 0.024 mg N₂O-N l^-1 [15].

In Sep 2014, the average farm excess-N₂ was 1.90 mg N₂-N l^-1 (max. 82 mg N₂-N l^-1; min. 0.0004 mg N₂-N l^-1) and the average drainage system was 0.076 mg N₂-N l^-1 (max, 0.18 mg N₂-N l^-1; min, 0.006 mg N₂-N l^-1 (S2 Fig, S2 Table). The up-gradient unit was characterised by emissions of 0.54 mg N₂-N l^-1. In the down-gradient area at shallow depth, excess-N₂ was elevated in well- to moderately-drained soils, with peak concentrations of 3.36 mg N₂-N l^-1 at 2.95–4.5 m bgl and 4.32 mg N₂-N l^-1 at 4.5–6 m bgl. Up-gradient imperfectly drained soils had elevated concentrations (4.12 mg N₂-N l^-1) at 6–9 m bgl and these were higher in central parts of the farm (6.52 mg N₂-N l^-1 and 6.82 mg N₂-N l^-1) at 11–13 and 16 m bgl, respectively (S2 Fig). Dissolved-N₂O averaged 0.03 mg N₂O-N l^-1 (max. 0.036; min. 0.0002 mg N₂O-N l^-1). The farm value for EF₅g(1) was 0.0039 mg N₂O-N/mg N input, compared with the IPCC default value of 0.0025 mg N₂O-N/mg N input for groundwater N₂O emissions. Both drainage and surface water emissions were below this default with only one in-field drain (D1, 0.0057 mg N₂O-N/mg N input) had a higher value (S3 Fig). The up-gradient unit was characterised by a high emission factor (0.0243 mg N₂O-N/mg N input), while the down-gradient unit (0–4.5 m bgl) had piezometers exceeding default values in the central area, reaching a maximum of 0.0081 mg N₂O-N/mg N on well to moderately drained soil. At intermediate depths (4.5–9 m bgl) values were above the default values towards the north (6, 0.0325 mg N₂O-N/mg N input) and south (23, 0.0114 mg N₂O-N/mg N input; 37, 0.0097 mg N₂O-N/mg N input). In contrast to excess-N₂, N₂O decreased with increasing depth, with almost no piezometer above the default values below 11 m bgl (S3 Fig).

A wide range of dissolved N₂O vs. total emissions (N₂O + N₂) was found, which suggests a variable rate of denitrification. If the denitrification rate is high keeping the NO₃^--N concentration below the contamination threshold, excess-N₂ is released due to completion of the process. Conversely, high dissolved-N₂O occurs where denitrification is limited [3]. Nitrification contributes to N₂O production with stable isotopes (δ¹⁸O and δ¹⁵N) of N₂O elucidating discrepancies in gas production and identifying N₂O sources [33–36]. N₂O-N was produced in situ and correlated with water table depth and k_s [13,15,21]. From Snider et al. [34] the expected range of δ¹⁸O-N₂O produced by nitrification-denitrification on site was 0–20‰ (S4 Fig), which includes 75% of the piezometers. Piezometers with a relative enrichment of δ¹⁸O-N₂O above these values are presumably only influenced by N₂O reduction via denitrification. In addition, using the δ¹⁸O-N₂O and δ¹⁵N-N₂O ranges reported by Li et al. [37] and in particular the δ¹⁵N-N₂O enrichment values, almost all locations have a N₂O signature characterised by relative enrichment in δ¹⁵N-N₂O, which reflects enrichment in the NH₄⁺ source due to its consumption by microbial processes.

Investigating connectivity

Stable isotope compositions for all water samples showed low spatial variability, with values between -6.4 and -3.4‰ for H₂O-δ¹⁸O and between -39.1 and -25.1‰ for H₂O-δD in September 2014. Similarly, values ranged between -6.2 and -4.0‰ for H₂O-δ¹⁸O and between -36.8 and -21.9‰ for H₂O-δD in June 2017 (Fig 4). To examine the groundwater interaction with the drainage system, water stable isotope values were compared with the Global Meteoric Water Line (GMWL, δD = 8*δ¹⁸O+10) [38] to infer their composition and provenance (Fig 4). Farm values had a lower slope than the GMWL. Water samples (from screened intervals of piezometers and boreholes) had stable isotopic values close to the GMWL. Some sub-surface locations had the same water isotope values as many groundwater boreholes and piezometers. However, for other locations, e.g. along the surface open drain, there was an enrichment in δ¹⁸O but not δD relative to the GMWL in 2014, while in both δ¹⁸O and δD in 2017 (Fig 4), therefore showing a second signature due to evaporation in the open drain. A third signature from the lake system exhibited enrichment in both δ¹⁸O and δD-H₂O (Fig 4).

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Fig 4. δ¹⁸O-H₂O versus δD-H₂O values for samples collected in Sep 2014 and June 2017 at the sampling locations.

Circles indicate groundwater and infield drain samples, squares indicate open ditch samples while triangles indicate samples from the lake system.

https://doi.org/10.1371/journal.pone.0219479.g004

Groundwater flow direction is from north to south, mirroring topography [14]. The isotopic composition of groundwater samples collected in Sep 2014 plots on a lower slope of the GMWL identified by Darling et al. [39], indicating a relative enrichment consistent with the high-humidity climate of the British Isles and findings of Gibson et al. [40]. Samples from the end-of-pipe and groundwater had the same signature, suggesting a common origin and interaction. However, other points in the open ditch were enriched in δ¹⁸O or in both δ¹⁸O and δD caused by migration and accentuated evaporation [41]. Overall, this indicates that water from a single source is connected along a groundwater, in-field pipe and open drain continuum.

Spatial variation in nitrate δ¹⁸O and δ¹⁵N in water samples

In 2014, the average δ¹⁵N composition of NO₃^- was 14.2‰ (6.2–54.9‰, Fig 5, S2 Table). The average δ¹⁸O composition of NO₃^- was 9.7‰ (2.3‰-28.2‰). Isotopic values within the drainage system varied between 5.86 and 7.86‰ for δ¹⁸O-NO₃^- and between 9.72 and 12.89 ‰ for δ¹⁵N-NO₃^-.

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Fig 5.

Top: scatterplot showing δ¹⁸O-NO₃^- vs. δ¹⁵N-NO₃^- for water samples collected in September 2014 superimposed onto δ¹⁸O and δ¹⁵N ranges for N-sources and processes by Yin et al. [49] and Baily et al. [14]. Bottom: scatterplot showing δ¹⁸O-NO₃^- vs. NO₃^--N values, identifying condition of inputs and denitrification rate. Whole circles identify wells with alternatively 1) NO₃^--N concentration <5.65 mg NO₃^--N l^-1 and high denitrification isotope signature (>10‰) i.e. exhibiting excess inputs of NO₃^--N that has been denitrified or 2) NO₃^--N >5.65 mg NO₃^--N l^-1 and low denitrification isotopic signature (<10‰) i.e. exhibiting contamination due to an insufficient rate of denitrification. Open circles identify wells that were discarded due to depth or where the NO₃^--N concentration was <5.65 mg NO₃^--N l^-1 in combination with a low denitrification isotopic signature (<10‰), i.e. exhibiting a situation of limited denitrification and low inputs.

https://doi.org/10.1371/journal.pone.0219479.g005

The up-gradient unit had isotopic values of 3.2 and 13.2‰ for δ¹⁸O-NO₃^- and δ¹⁵N-NO₃^-, respectively. In the down-gradient unit at 0–4.5 m bgl, locations showed higher enrichment in the central and south area. This range was soil drainage class led e.g. moderate class at location 20 (17.1 and 25.6‰) and poor class at location 26 (17.6 and 52.7‰) for δ¹⁸O-NO₃^- and δ¹⁵N-NO₃^-, respectively. The well drained area around location 21 showed enrichment only for δ¹⁵N-NO₃^- (54.9‰). Enriched values occur between 4.5–6 m bgl at locations 24 and 25 (poor drainage) and near locations 35 and 36 (imperfect drainage). Deeper screen intervals between 6–9 m bgl, show an enrichment for δ¹⁵N-NO₃^- in the north (11, 28.2 and 13.0‰ for δ¹⁸O-NO₃^- and δ¹⁵N-NO₃^-) and south areas. Similar enrichments occur between 11–13 and 13–16 m bgl for both δ¹⁸O-NO₃^- and δ¹⁵N-NO₃^- at location 13 (19.5 and 21.9‰ for δ¹⁸O-NO₃^- and δ¹⁵N-NO₃^-) and for δ¹⁵N-NO₃^- at location 38 (20.9‰).

Importantly, comparing these values with [14] for the same wells, 75% of δ¹⁵N-NO₃^- and δ¹⁸O-NO₃^- values within 3‰ of those in 2008. In 2017, 44% and 31% of the wells were within 3‰ of those in 2008 for δ¹⁵N-NO₃^- and δ¹⁸O-NO₃^- respectively while 38% and 41% of the wells were within 3‰ of the 2014 values for δ¹⁵N-NO₃^- and δ¹⁸O-NO₃^- respectively. These three sampling events (2008, 2014 and 2017) confirm the findings of [14] that there is a consistent temporal signal.

In 2014, NO₃^--N was spatially variable with some locations exceeding 5.65 mg l^-1 (Fig 2, S2 Table). In 2008, three main areas: the north end (surrounding 5, well-drained soil), with a value of 9.5 mg NO₃^--N l^-1 (±2.9); the central area (surrounding 14, well-drained soil), with a value of 10.3 mg NO₃^--N l^-1 (±4.5); and south end (surrounding 28, well- to moderately- drained soil), with a value of 7.3 mg NO₃^--N l^-1 (±2.4) [14] were elevated. High NO₃^--N was attributed to old DSW irrigation areas and farmyard leachate (see [14]). In Sep 2014 NO₃^--N < 5.65 was evident for all locations (Fig 5). In terms of N source groundwater NO₃^--N isotopic composition clustered within the manure/sewage value range and along a 1:1–1:2 slope, suggesting a common organic source for these, and denitrification as the main biotransformation process (Fig 5) [41–43]. The denitrification lines for 2014 and 2017 samples share the same trend, which points to denitrification as the main process leading N attenuation. Distinguishing between denitrification and DNRA is difficult as the isotope effect of DNRA has still not been investigated [44,45].

The use of a modified Rayleigh equation to estimate NO₃^--N attenuation [46] is acceptable as the source on site is uniform and the source of water on site has been established. Values between -3 to -30‰ have been reported for ε_denit [42,47]. Herein, the ratio of δ¹⁸O:δ¹⁵N enrichment during denitrification remained constant. Therefore, enrichment in both δ¹⁵N-NO₃^- and δ¹⁸O-NO₃^- can be attributed to biotransformation of NO₃^--N and is directly proportional to the degree of denitrification (f) [45,46]. Across the entire site, only a few wells showed high temporal variability, possibly due to management, compared with values in 2008 and 2014–2017. Such wells showed alternatively high values of δ¹⁸O-NO₃^- or δ¹⁵N-NO₃^- and exhibited a shift from the 1:1–1:2 slope (Fig 5) [48]. From this group two other signatures emerged: a high enrichment of δ¹⁵N-NO₃^- possibly from surface NH₃ volatilization, and high δ¹⁸O-NO₃^- values, possibly due to an atmospheric source or synthetic fertilizer as nitrate source (Fig 5).

The subset of data points, close to the 1:1 isotope ratio line, where denitrification was the dominant process, was further examined by eliminating points with a fertiliser and ammonia volatilisation signature (Fig 5). A plot of δ¹⁸O-NO₃^- versus NO₃^--N revealed three main groups, each identifying a specific condition of inputs and extent of denitrification (Fig 5). The first group had a low NO₃^--N concentration (<5.65 mg NO₃^--N l^-1) associated with limited denitrification isotopic signature (<10‰) and low inputs. The second group had a low NO₃^--N concentration but high denitrification isotope signature (>10‰), indicating that excess inputs of NO₃^--N have been denitrified. The final group had high NO₃^--N (>5.65 mg NO₃^--N l^-1) and low denitrification isotopic signature (<10‰), highlighting contamination but limited denitrification. The first group was discarded from analysis as not influential in terms of denitrification and contamination, while groups with NO₃^--N above 5.65 mg NO₃^--N l^-1 or δ¹⁸O-NO₃^- above 10‰ were selected. Poorly-drained soils are characterised by high denitrification potential and low NO₃^--N, whereas well-drained soils had low denitrification potential and higher NO₃^--N. These isotopic enrichment patterns followed soil permeability patterns which were consistent with the findings of Fenton et al. [13] (Fig 5).

Natural isotopic abundance for NH₄⁺-N was measured in wells with detectable NH₄⁺-N concentration. The average δ¹⁵N-NH₄⁺ farm concentration was 18.5‰ (max 34.3‰ corresponded with breach of MAC; min 3.3‰). The site average N₂O stable isotope composition was 14.2‰ for δ¹⁸O-N₂O and -17.1‰ for δ¹⁵N-N₂O, respectively. Maximum values were 46.7‰ and 21.4‰ for δ¹⁸O-N₂O and δ¹⁵N-N₂O, respectively, while minimum values were -14.5‰ and -32.5‰. NH₄⁺-N polluted wells occur mainly in the central area of the farm (Fig 3) with moderate- to imperfectly-drained soil. A high NH₄⁺-N concentration was evident in wells showing relative enrichment in δ¹⁸O-NO₃^- and with a synthetic fertilizer source signature (Fig 5). However, a high NH₄⁺-N concentration could occur without this signature, due to in-situ transformation of NO₃^- to NH₄⁺ by DNRA [11,15]. Even though denitrification and DNRA occur under similar environmental conditions, DNRA has rarely been observed with respect to denitrification but is an important process under anaerobic conditions with high (>12) C/NO₃^- ratio [49,50]. High NH₄⁺-N location showed lower (p<0.05) dissolved oxygen concentration (av. 0.65 mg l^-1) and NO₃^--N (av. 0.94 mg NO₃^--N l^-1) than low NH₄⁺-N locations (av. 3.72 mg l^-1 and 3.46 respectively mg NO₃^--N l^-1). No significant differences between high and low NH₄⁺-N location were found in terms of DOC concentrations or groundwater level. Since no direct fractionation factors exist for DNRA [51], δ¹⁵N-NH₄⁺ values were measured to assess a possible role of DNRA. As per Rayleigh fractionation, we hypothesised that DNRA would lead to a δ¹⁵N-NH₄⁺ value significantly less enriched than at a location where denitrification was the dominant process. With only four locations having a C/NO₃^- ratio >12, δ¹⁵N-NH₄⁺ signatures did not show any distinct pattern. This could indicate that DNRA was of secondary importance and restricted to a few locations at high depth or micropores within the soil profile (S5 Fig).

Conceptual diagram of the site to inform soil nitrogen attenuation potential

Integrating all data (Fig 6), four main location groups were distinguished with respect to potential for soil N attenuation (Fig 1, S2 Table):

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Fig 6. Conceptual diagram of the two-tiered system beneath the site: 1) a shallow migration pathway in poorly-imperfectly drained soils with high N_O³-N attenuation which is not disrupted by the artificial drainage system to the outlet; 2) a deep migration pathway under moderately-well drained conditions where N_O³-N attenuation is lower therefore leading to its transformation in NH₄⁺-N.

https://doi.org/10.1371/journal.pone.0219479.g006

Group 1 (high N attenuation potential): locations with NH₄⁺-N and NO₃^--N concentrations below MAC. Drainage class of these soils are imperfectly to moderately-well drained and drainage water from these areas does not represent a threat to water quality. This condition reflects relatively high NO₃^--N attenuation rates, via denitrification. Generally low N₂O emissions inferred completeness of denitrification, although some locally elevated N₂O production suggested concurrent secondary processes (e.g. nitrification), with a possible threat to air quality.

Group 2 (moderate N attenuation potential): locations with a NO₃^--N concentration below the contamination threshold, but with high N₂O production. Drainage classes are well to moderately-drained and drainage water from these soils does not represent a threat to water quality, but instead is a problem in terms of GHG emissions. Such emissions are due to a low capacity for denitrification, resulting in low NO₃^--N, possibly coupled with additional N₂O emissions from nitrification.

Group 3a (low N attenuation potential, DSW irrigation): locations with a NO₃^--N concentration above significant contamination. Drainage classes are mainly well to moderately-well drained, with a higher permeability than those in Group 1 and drainage water from these locations represents water and air quality issues (due to high N₂O emissions, which indicate incomplete denitrification). This results from a medium-low potential for NO₃^--N attenuation by denitrification and if drained will present a threat to groundwater and GHG emissions.

Group 3b (low N attenuation potential): locations with low NO₃^--N but NH₄⁺-N above NH₄⁺-N MAC and high N₂O emissions. Drainage classes are well to imperfectly-drained with generally low potential for denitrification and drainage water is a threat to groundwater and GHG emissions.

De Klein et al. [52] observed that more than 50% of EU dairy farms have high N surpluses (>200 kg N ha^-1), with only 7% showing values below 100 kg N ha^-1. The N surplus on our Irish farm was high and above the national average of 175 kg N ha^-1 [53]. As a direct continuum has been shown to exist N can transported and transformed as follows (Fig 6):

1. a shallow migration pathway in poorly- or imperfectly-drained soils with high water purification capability. The drainage system does not disrupt this capacity but instead conveys clean water to the exit point of the farm;
2. a deep migration pathway in moderately- and well-drained soils where the water purification capability is lower. This facilitates leaching of N, which is then converted at depth to NH₄⁺-N and migrates off site along deeper groundwater pathways.

Development of farm management strategy for nitrogen

The farm is operating within the current regulations [54–56], following the EU Nitrates Directive [57], but these assume homogeneous soil-subsoil-geology without variation in N natural attenuation capacity. The present study shows that four attenuation capacity areas exist, with Group 3a and 3b areas representing the most vulnerable. N loads need to be diverted away from these areas e.g. DSW produced needs to be reduced and irrigation on the farm should avoid group 3a and b areas, and be reduced on group 2 areas.

Historically, the DSW irrigation system has been moved around (e.g. after a long period in 2005 it was moved from heavier soils along the central drainage ditch to avoid overland flow losses) and since then it is positioned behind the farmyard on well- to moderately-drained soils. This static position has concentrated a high N load in this area and can be correlated with elevated NO₃^--N in down-gradient groundwater. Presently all DSW is collected in a three-chamber separation farmyard underground tank (total capacity ~55,000 l) with a variable retention time depending on volume of washings and rainfall. In periods where land is saturated and spreading with Roto-Rainer is not possible, DSW can be stored in an earthen bank lagoon. This DSW is spread with an umbilical cord system within the downstream farm on the three main sections below the built-up area in Fig 1 (plots with locations 28, 33 and 35).

Farmyard measures must be optimised to manage the volume of DSW produced on the farm, through 1) Reduction of DSW production—Diversion of all clean water to a clean water outfall, preventing clean water from becoming soiled and restructuring of the yard to minimise DSW generation; 2) Optimisation of DSW spreading—Use of the information gained from the present study (heterogeneous soil functionalities) to better use the entire farm to spread DSW. The following management options are considered to be viable in terms of reducing the N load:

1. Short term management options: 1) Moving DSW irrigation to high N attenuation areas (Group 1 imperfectly- to moderately-well drained soil locations), or moderate N attenuation areas (Group 2 well- to moderately-drained soil locations), where no overland flow component could affect the open ditch network; 2) Removal of the N load from vulnerable areas that will allow the N biogeochemical time lag to diminish over time, which should result in a decreased groundwater NO₃^--N concentration.
2. Long term management option: minimise the N source whilst using the lagoon for storage. This provides flexibility for DSW to be 1) land spread when its fertilizer replacement value is high, or 2) discharged to a constructed wetland installed at the end of the farm (Group 1 location) at other times, e.g. over winter. The lagoon concept has been considered on dairy farms [58,59]. The loss of land for the lagoon would not be severe as full utilisation of the area currently being used for DSW application is difficult at times due to ground trafficability and grass palatability. In addition, the labour lost through the management of an irrigation system would be gained for other activities on the farm. Significant energy saving is also achievable with the significant reduction in pumping of DSW. One consideration here is to design the system for expansion in the event that the discharge licence is not met over time, in which case a de-sludging specification should be considered.