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Environmental Science and Pollution Research

, Volume 26, Issue 3, pp 2089–2104 | Cite as

Evaluation of sources and fate of nitrates in the western Po plain groundwater (Italy) using nitrogen and boron isotopes

  • Manuela LasagnaEmail author
  • Domenico Antonio De Luca
Groundwater under threat from diffuse contaminants: improving on-site sanitation, agriculture and water supply practices

Abstract

Diffuse nitrate pollution in groundwater is currently considered one of the major causes of water quality degradation. Determining the sources of nitrate contamination is an important first step for a better management of water quality. Thus, the isotopic composition of nitrate (δ15NNO3 and δ18ONO3) and boron (δ11B) were used to evaluate nitrate contamination sources and to identify geochemical processes occurring in the shallow and deep aquifers of the Turin-Cuneo plain (NW Italy). The study area is essentially an agricultural zone, where use of synthetic nitrogenous fertilizers and organic manure is a common practice and the connection to sewer services is locally lacking. Also livestock farming are highly developed. A groundwater sampling campaign was performed on 34 wells in the shallow aquifer and 8 wells in the deep aquifers, to analyze nitrate, chloride, boron, δ15NNO3, δ18ONO3 and δ11B. Isotope data of nitrate indicate that nitrate contamination in the Turin-Cuneo plain originates from mixtures of synthetic and organic sources, slightly affected by denitrification, and manure or septic tank effluents. Moreover, boron isotopes were used to discriminate further among the main anthropogenic sources of pollution. The analyses results confirm that both animal manure and domestic sewage, especially under the city of Turin, can contribute to the nitrate contamination. The isotope analysis was also used for the evaluation of denitrification and nitrification processes: contrary to expectations, a significant denitrification phenomenon was assessed only in the shallow unconfined aquifer, especially in the Poirino Plateau, the most contaminated sector of the study area.

Keywords

Nitrogen stable isotopes Boron Nitrate sources Groundwater pollution NW Italy 

Introduction

Nitrate pollution in groundwater is currently one of the major cause of water quality degradation. Starting from the second half of the past century, the global increase in the use of N-fertilizers (synthetic and organic) has led to increased nitrate leaching and runoff. Consequently, large regions of the world suffered from nitrate contamination of surface and groundwater, especially in agricultural areas. The correlation between agriculture and nitrate concentrations in groundwater was observed in many studies coming from different countries (Agrawal et al. 1999; Nolan and Stoner 2000; Harter et al. 2002; Debernardi et al. 2008; Capri et al. 2009; Petitta et al. 2009; Sacchi et al. 2013). The extensive use of fertilizers for agricultural purposes, in fact, is considered to be the main non-point source of nitrate contamination in groundwater (Liao et al. 2012; Postma et al. 1991; Baker 1992; Chowdary et al. 2005). However, nitrate pollution in groundwater can also be caused by other sources, such as septic systems (Al-Agha 1999; MacQuarrie et al. 2001).

The Nitrates Directive (European Commission 1991) is one of the earliest pieces of EU legislation aimed to protect water quality across Europe by preventing nitrates from agricultural sources polluting ground and surface waters and by promoting the use of good farming practices. One of the adopted action was the Designation of “Nitrate Vulnerable Zones” (NVZs). NVZs are territories which drain into polluted waters or waters at risk of pollution and which contribute to nitrate pollution. As regards water resources, groundwater containing or that could contain (if no action is taken to reverse the trend) more than 50 mg/L of nitrates are defined as polluted water, or at risk of pollution.

In Piedmont, that represents the westernmost region in Italy, NVZs correspond to the 37.6% on utilized agricultural area and 54% on regional plain areas (Ministry for Environment et al. 2010).

In the NVZs action programs must be implemented by farmers on a compulsory basis, including limitation of fertilizer application (mineral and organic) and other measures at farm scale. The application of these measures can have negative impact on farm economics, especially for small farms (ADAS 2011). Consequently, the definition of the different nitrate sources in groundwater (sewage, animal manure, chemical fertilizer, natural soil mineralization) has major importance, especially for the stakeholders that must implement management actions for groundwater protection.

To define the sources of nitrates in groundwater, the stable nitrogen (δ15NNO3) and oxygen (δ18ONO3) isotope of nitrates have been extensively used (Aravena et al. 1993; Seiler 2005; Sacchi et al. 2007; Re and Sacchi 2017).

Nitrates from mineral fertilizers are characterized by the atmospheric values of δ15NNO3 (approximately 0‰) and δ18ONO3 (approximately 23.5‰), whereas nitrates derived from organic fertilizers, livestock effluents and sewage are usually enriched in 15N (δ15NNO3 varies from +8 to +20‰) because of ammonium volatilization (Kendall 1998). The correlation between the δ15NNO3 and δ18ONO3 isotopic indicators can also provide data on the nitrification and denitrification in an area. Aravena and Robertson (1998) demonstrated a concurrent increase in the δ15NNO3 and δ18ONO3 of groundwater nitrates during denitrification. During nitrification, the δ18ONO3 ratio changes to lower values because of the incorporation or the exchange with δ18OH2O of groundwater. The typical range of δ18ONO3 for nitrified synthetic fertilizers generally is 5‰ to 15‰; it also depends on the groundwater isotopic composition (Sacchi et al. 2013). Furthermore, nitrification process produces a positive correlation between the nitrate concentrations and δ15N of the nitrates (Choi et al. 2003). Thus isotopic composition of nitrate can help to discriminate between synthetic fertilizers, mineralized synthetic fertilizers, manure/septic system effluents and soil organic matter as nitrate sources in groundwater. However, the use of these isotopes is not useful to further discriminate between manure and septic system effluents that display the same nitrate isotopic composition.

In recent studies boron isotopes (δ11B) were used coupled with δ15N of the nitrates to discriminate between synthetic fertilizers, sewage and septic plumes, cattle, poultry and hog manure sources (Widory et al. 2004, 2005; Seiler 2005). Boron (B) has generally a ubiquitous distribution in groundwater, because it commonly exists as a minor constituent (Vengosh 1998). Naturally occurring B is mostly derived from weathering of rocks; thus δ11B composition can be used to identify pollution sources if the contaminants have δ11B values distinct from the δ11B composition in groundwater (Vengosh et al. 1994; Bassett et al. 1995). More specifically, detergents coming from evaporites, synthetic fertilizers, manure and sewage have high content of B and different δ11B signature (Vengosh et al. 1994; Komor 1997; Widory et al. 2005).

Komor (1997) and Bassett et al. (1995) first used B isotopes as comigrating tracers of NO3, but only Komor (1997) and Widory et al. (2004) referred the B isotope composition of input from agriculture (e.g., hog manure, cattle feedlot runoff, synthetic fertilizers) and combine N and B isotopes in order to distinguish between different NO3 sources in groundwater and surface water.

Microbial processes and groundwater redox that govern N isotope evolution (denitrification/nitrification) have no consequence on B (Bassett et al. 1995). The key factors that affect the B isotopic composition in groundwater are the B sources and adsorption to clays; only in extremely saline environments, mineral precipitation affects the B isotopic composition (Seiler 2005).

A recent study (Pennisi et al. 2013) compiled a database of isotopic data on potential boron contaminants of natural waters (ISOBORDAT database). Their isotopic composition, which spans over a wide range, helps to identify boron contamination sources and geochemical processes in water.

The purpose of this paper is to define the nitrate contamination sources in the shallow and deep aquifers of Turin-Cuneo plain (NW Italy) using isotopic composition of NO315NNO3 and δ18ONO3) and boron (δ11B). Furthermore, the investigation aims to identify the processes affecting nitrate concentrations in groundwater, such as nitrification and/or denitrification.

Study area

The study area is located in Piedmont (northwestern Italy) and corresponds to the plain between Turin and Cuneo cities (Turin-Cuneo plain). The study area can be further distinguished in different sub-areas: the Po Plain area in Turin Province (Turin plain), a plain area in Cuneo Province (Cuneo Plain) and a plateau area in Turin and Asti Province (Poirino Plateau). The minimum altitude is 200 m above sea level (a.s.l.) in the eastern sector of the plain, corresponding to the confluence of the Stura di Demonte River and Tanaro River, and the maximum altitude is 600 m a.s.l. in the southern sector, near Cuneo town. The plain area is underlain by an important groundwater resource due to its size, the characteristics of sediments and the relatively high rate of recharge in the region (Lasagna et al. 2016a).

Hydrogeological setting and land use

The study area is featured by superimposed hydrogeological complexes (Fig. 1): the pre-Pliocene complex, the Pliocene marine complex, the Villafranchian transitional complex, the Quaternary deposits complex (Bortolami et al. 1976; Bove et al. 2005). A simplified cross-section was added, to clarify the hydrogeological setting in the alluvial Po plain (Fig. 2).
Fig. 1

Hydrogeological map of Turin-Cuneo plain. In the figure the hydrogeological complexes of the study area and the piezometric map of the shallow aquifer are reported. The black line in Turin plain represents the line of the cross-section of Fig. 2

Fig. 2

Simplified cross-section of the Turin Po plain, with the indication of the main hydrogeological complexes

The Turin Hill and the alpine basement border the plain on the west and the east, respectively. The Turin Hill is represented by the Tertiary Piedmont Basin (TPB) deposits, mainly sandy clay, marl, highly consolidated (b in Fig. 1). These sediments have a low permeability, being only locally permeable by fissuration, and do not contain any significant aquifers.

The alpine basement mainly consists of crystalline rocks (a in Fig. 1), mostly impermeable or slightly permeable by fissuration, and subordinately by calcareous rocks (a’ in Fig. 1), permeable for karst.

The shallow unconfined aquifer is hosted in the Quaternary alluvial deposits complex (Middle Pleistocene-Holocene), consisting of coarse gravel and sand of fluvial or fluvio-glacial origin, with subordinate silty-clayey intercalations (g in Fig. 1). It generally shows a high permeability (k = 5*10−3÷5*10−4 m/s) and represents an important aquifer whose water table is directly connected to surface drainage in the region. Grain size normally decreases from mountains to low plain, in correspondence to the Po River. Its thickness ranges between 20 and 50 m; in spite of the variable thickness of the aquifer, it has a high productivity and has regional importance. The presence of thick and relatively continuous layers of silt or clay-rich deposits generally marked the base of the shallow aquifer (Canavese et al. 2004; Bove et al. 2005). Groundwater temperature values mostly range between 12 and 14 °C in rural areas shallow aquifer (Barbero et al. 2016; Bucci et al. 2017).

In the Poirino Plateau, located on the eastern side of the Turin-Cuneo Plain, the Quaternary alluvial deposits complex consists of silt and clay, with rare sandy–gravelly levels (f in Fig. 1); it has a thickness between 10 and 30 m. In this area, the shallow aquifer has small thickness and low productivity.

The deep aquifers are located in the Villafranchian transitional complex and in the Pliocene marine complex. The first is a multilayer aquifer, and it consists of alternating clayey silt, sand and small gravel (Middle Pliocene-Lower Pleistocene), thus creating an alteration of permeable and semi-permeable layers (e in Fig. 1).

The Pliocene marine complex (Lower-Middle Pliocene) consists of two different units: the Asti sands (d in Fig. 1), permeable and hosting important confined aquifers, and the Lugagnano clay (c in Fig. 1), typically an aquitard with low permeability.

In the Turin-Cuneo plain, the shallow aquifer is mainly supplied by rainfall and rivers at the outlet of the valleys on the plain. The general flow direction in the shallow aquifer is SW-NE in the southern part of the plain, and S-N in the northern part (Lasagna et al. 2016a). In the Poirino Plateau, the groundwater generally flows towards the W, i.e., towards the Po River, which represents the main watercourse of the study area.

The hydraulic gradient of the shallow aquifer varies, in the northern sector of the Turin-Cuneo plain (Turin Plain), between 3%, e.g., at the edge of the Alps and 0.1% in the low plain. Along the transitional zone from the higher to lower plain, the hydraulic gradient decrease from 0.6% to inferior than 0.3% values; this sector, in which typical lowland springs (fontanili) emerge (De Luca et al. 2009, 2014), is called fontanili line. In the centre of the Turin-Cuneo plain, the hydraulic gradient normally ranges between 0.01% in the central sector and 0.25% near the Alps. On the south of the Cuneo plain, the hydraulic gradient is high near the Alps (0.2%) and decreases to 0.02% towards the central plain.

The depth to groundwater in shallow aquifer is very variable: on the low plain and near the rivers, the water table is generally less than 5 m deep, whereas it ranges depths between 20 and 50 m close to the Alps. On the fontanili line, the depth to groundwater varies from 1 to 3 m. On the Poirino Plateau, the groundwater depth is generally low, varying between 0 m and 5 m.

In the villafranchian transitional complex and in the Pliocene marine complex (Asti sand), confined and semiconfined aquifers are generally hosted, with a flow direction generally similar to the shallow aquifer. Only locally, e.g. in the Poirino Plateau, the flow directions are very different (Lasagna et al. 2014).

The land use is conditioned by the anthropic activities that in the area are represented by agriculture (mainly cereals and forages cultivations) and farms (especially cows and pigs) (Regione Piemonte 2008). In the small towns domestic waste water is locally not connected to sewerage.

Moreover, industrial areas are mainly located in the peripheral areas of Turin while quarrying areas are located near the main streams, especially the Po river, for the extraction of sand and gravel (De Luca et al. 2007; Castagna et al. 2015a, b).

As regards irrigation, the Po river basin is very rich of water, which is used for various purposes, including irrigation. A complex network of canals allows water transfer used for irrigation from the Po river and its affluents at the foothills of the Alps downwards, to the productive plains of the Po river basin. The estimated average efficiency of transport and distribution of the network is approximately 69% (Ministry for Environment et al. 2010; Comina et al. 2014; Perotti et al. 2015; Clemente et al. 2015). Irrigation methods are different in the different agricultural area. In Piedmont surface irrigation or furrow irrigation are the prevailing techniques.

The land use in the plain can be identified as the cause of a groundwater nitrate contamination, especially for the shallow aquifer (Lasagna et al. 2005, 2013; Debernardi et al. 2008).

Designation of nitrates vulnerable zones (NVZs)

The designation of nitrates vulnerable zones (NVZs) in Italy falls under the competence of the Regions. The NVZs designation in Piedmont took place in different steps, starting from 2002 and according to the Nitrates Directive (European Commission 1991).

The first designation is dated back to 2002 (DPGR 9/R 2002). Designation is based on the criteria of article 3 and Annex 1 of Nitrate Directive, more specifically areas of land which drain into polluted waters or waters at risk of pollution and which contribute to nitrate pollution were designated as NVZs. In Piedmont NVZs comprised both the areas already contaminated, with nitrate concentrations in groundwater equal or higher than 50 mg/L on the basis of the results of monitoring programs in 2000–2001, and conservatively the areas with nitrate concentrations equal or higher than 40 mg/L. The areas identified by the first designation correspond to 220,121 ha and represent about 8% of territory, approximately 30% of the plain.

Water Defence Plan (PTA), drawn up by the Piedmont Region in 2004 (DGR 23–13,437 2004) and 2006 (DGR. 28–2845 2006), included in the NVZs also the flood zones A and B of rivers, according to the definition in PAI (Piano stralcio per l’Assetto Idrogeologico del Fiume Po - Po river basin plan for hydrological management) approved by Autorità di bacino del fiume Po (AdBPo 2001).

More specifically Zone A is ordinary floods area for a return time up to 50 years and Zone B is flood area for a return period up to 200 years; it corresponds to the area confined by the embankment system. In this way further 54,162 ha of territory were designated as NVZs, in order to prevent any phenomenon of run-off and to contain the transport of pollutants, including nitrogen, towards the surface water body.

In 2007 with a regional regulation (DPGR 12/R 2007) were designated additional territories, potentially vulnerable to nitrates from agricultural sources. The vulnerability has been assessed on the basis of a scientific study that took into account some agro-environmental parameters: (a) intrinsic vulnerability of shallow aquifer in Piedmont plain, evaluated by Earth Science Department of Turin University (Bove et al. 2005); (b) protective capacity of the soil, assessed by IPLA (Institute for Wood Plants and Environment); (c) the N-surplus, defined as the algebraic value obtained by the difference between the N-inputs to the soil (organic nitrogen and synthetic N-fertilizers) and the N-output with crop products, evaluated by Agroselviter Department of Turin University (Sacco et al. 2006). In 2007 further 128,094 ha were designated as NVZs.

At last in 2013 the delimitation of Zones A and B were updated and consequently new territories were designated as NVZs.

Currently in Piedmont NVZs occupy 407,500 ha that represent approximately 38% of the regional utilized agricultural area and about 54% of the plain area (Fig. 3).
Fig. 3

Nitrate concentrations in the shallow and deep aquifers (October 2014). In the pie chart, the distribution of nitrate concentrations in shallow aquifer is reported. The number represent the denomination of water samples

Methods

A groundwater monitoring campaign was performed in October 2014 on 42 water wells and monitoring wells in Turin-Cuneo plain. It includes 34 sampling points in the shallow aquifer (Table 1) and 8 in the deep aquifers. The most of the sampling points in the shallow aquifer are referred to monitoring wells of the Monitoring Network of the Piedmont Region, managed by the Regional Agency for the Protection of the Environment (ARPA Piemonte). These sampling points are located as follows: 17 in Cuneo plain, 11 in Turin plain and 6 in Poirino Plateau.
Table 1

Groundwater sampling points in the shallow and deep aquifers in October 2014 monitoring campaign and results of chemical and isotopic analyses. The detection limits are 15 mg/L for boron is and 0.05 mg/L for nitrite

Sample name

Area

Town

Typology

Use

Depth (m)

EC (μS/cm)

NO3 (mg/L)

NO2 (mg/L)

Cl (mg/L)

B (μg/L)

δ15NNO3

δ18ONO3

δ11B

Shallow aquifer

 001

Turin plain

Cumiana

Well

Domestic

40

468

29.7

< 0.05

47.4

< 15.0

13.81

10.5

 002

Turin plain

Airasca

Well

Domestic

20

335

40.7

< 0.05

13.9

< 15.0

6.46

8.8

 003

Turin plain

Piscina

Well

Domestic

30

350

21.7

< 0.05

22.3

< 15.0

5.47

7.5

 005

Turin plain

Osasco

Well

Agricultural

13

319

22.7

< 0.05

5.2

< 15.0

6.20

7.7

 006

Cuneo plain

Fossano

Monitoring well

Monitoring

20

518

29.5

< 0.05

10.0

< 15.0

4.75

6.5

 007

Cuneo plain

Fossano

Monitoring well

Monitoring

30

752

58.5

< 0.05

8.0

< 15.0

7.30

5.1

 008

Cuneo plain

Tarantasca

Monitoring well

Monitoring

54.5

689

27.4

< 0.05

5.9

< 15.0

6.19

6.4

 009

Cuneo plain

Busca

Monitoring well

Monitoring

75

570

16.4

< 0.05

4.1

< 15.0

4.90

7.1

 010

Poirino Plateau

Poirino

Well

Agricultural

40

612

25.5

< 0.05

19.2

< 15.0

7.55

10.7

 011

Cuneo plain

Monasterolo di Savigliano

Well

Drinking water

54

735

30.4

< 0.05

9.5

< 15.0

8.68

7.5

 012

Cuneo plain

Savigliano

Well

Drinking water

70

664

27.5

< 0.05

8.6

< 15.0

8.15

8.0

 013

Cuneo plain

Savigliano

Well

Drinking water

60

683

30.2

< 0.05

6.4

< 15.0

7.30

7.1

 014

Poirino Plateau

Poirino

Well

Agricultural

14

622

27.3

< 0.05

25.8

56.6

17.54

13.6

18.05

 017

Poirino Plateau

Riva Presso Chieri

Well

Domestic

17

880

28.7

< 0.05

38.1

86.1

13.53

10.0

16.77

 018

Poirino Plateau

Riva Presso Chieri

Well

Agricultural

13

678

46.0

< 0.05

30.5

< 15.0

16.58

11.9

 019

Turin plain

Carmagnola

Well

Agricultural

28

566

73.2

< 0.05

23.0

< 15.0

7.29

8.3

 020

Cuneo plain

Fossano

Well

Agricultural

44

755

40.1

1.02

13.6

< 15.0

8.47

7.1

 021

Turin plain

Carignano

Well

Domestic

15

603

5.3

< 0.05

9.1

< 15.0

6.56

14.9

 022

Cuneo plain

Rocca de’ Baldi

Well

Agricultural

9

404

34.4

< 0.05

13.4

< 15.0

15.61

10.0

 023

Cuneo plain

Narzole

Well

Agricultural

8

559

45.2

< 0.05

21.8

< 15.0

9.30

8.6

 024

Cuneo plain

Piozzo

Well

Domestic

8

465

26.7

< 0.05

18.0

< 15.0

7.36

6.1

 026

Turin plain

Villafranca Piemonte

Well

Drinking water

60

222

10.4

< 0.05

3.7

< 15.0

5.69

8.3

 027

Cuneo plain

Bra

Monitoring well

Monitoring

12

474

22.5

< 0.05

17.3

< 15.0

7.35

7.4

 028

Cuneo plain

Racconigi

Monitoring well

Monitoring

25

866

56.4

< 0.05

26.5

< 15.0

8.59

7.0

 029

Cuneo plain

Caramagna Piemonte

Monitoring well

Monitoring

15

626

54.8

1.13

28.1

< 15.0

8.41

8.3

 030

Cuneo plain

Scarnafigi

Monitoring well

Monitoring

25

440

19.9

< 0.05

6.5

< 15.0

7.49

6.2

 031

Cuneo plain

Cuneo

Monitoring well

Monitoring

60

498

12.0

< 0.05

5.5

< 15.0

4.00

8.4

 032

Poirino Plateau

Villanova d’Asti

Well

Agricultural

9

1355

177.5

< 0.05

101.2

< 15.0

20.66

11.8

 033

Poirino Plateau

Villanova d’Asti

Well

Agricultural

11

940

106.6

< 0.05

80.1

< 15.0

18.42

12.2

 034

Turin plain

Scalenghe

Monitoring well

Monitoring

25

558

50.6

< 0.05

14.1

< 15.0

7.37

6.4

 035

Turin plain

Lombriasco

Well

Drinking water

66

530

11.0

< 0.05

8.6

< 15.0

7.21

8.2

 037

Cuneo plain

Trinità

Well

Domestic

22

437

58.6

< 0.05

14.7

< 15.0

10.49

4.2

 041

Turin plain

Torino

Monitoring well

Monitoring

42.2

775

35.8

< 0.05

27.6

52.80

7.94

4.5

8.37

 042

Turin plain

Candiolo

Monitoring well

Monitoring

20

519

41.1

< 0.05

20.9

20.20

7.11

4.6

12.99

Deep aquifers

 004

Turin plain

Scalenghe

Well

Drinking water

80

608

37.7

< 0.05

12.3

< 15.0

7.29

9.3

 015

Cuneo plain

Cavallermaggiore

Well

Drinking water

103

580

26.7

< 0.05

11.1

< 15.0

8.70

8.2

 016

Cuneo plain

Genola

Well

Drinking water

65

576

20.0

< 0.05

5.9

< 15.0

6.61

7.1

 025

Poirino Plateau

Poirino

Well

Agricultural

50

492

5.4

< 0.05

12.0

< 15.0

23.6

 036

Cuneo plain

Fossano

Well

Drinking water

94

515

26.3

< 0.05

12.2

< 15.0

9.62

6.0

 038

Cuneo plain

Bene Vagienna

Well

Drinking water

150

407

7.6

< 0.05

5.7

< 15.0

6.05

21.6

 039

Cuneo plain

Sanfrè

Well

Drinking water

107

472

40.2

< 0.05

15.2

< 15.0

6.68

5.8

 040

Cuneo plain

Trinità

Well

Drinking water

114

447

24.0

< 0.05

46.2

< 15.0

9.84

6.7

All the sampling points of the deep aquifers are wells, also belonging to the Monitoring Network of the Piedmont Region, mostly used for drinking water purposes: 6 are located in Cuneo plain, 1 in Turin plain and 1 in Poirino Plateau.

The groundwater was sampled with a submersible pump, after the purge of the well/monitoring well. The purge time was determined based on the stabilization of water-quality indicator parameters included pH, electrical conductivity (EC), dissolved oxygen concentration and temperature. A Hydrolab Quanta G multiprobe meter was used to analyze the water-quality indicator parameters.

The groundwater collection was done using four polyethylene bottles for the different analyses: 0.5 L for nitrate, nitrite and chloride composition analyses; 0.5 L for boron concentration analyses; 0.5 L for boron isotopes analyses; 2 L for nitrate isotopes analyses. After the sampling, the water samples for nitrate isotopes analyses were acidified to pH 3 with concentrated HCl.

Groundwater samples were analyzed for EC, nitrate, nitrite, chloride and boron content and isotopic composition of NO315NNO3 and δ18ONO3) and boron (δ11B).

The EC was measured immediately after sampling in the field with a portable conductivity meter HANNA, with a resolution of 0.1 μS/cm and an accuracy within 1% of the reading.

The chemical analyses of nitrate, nitrite and chloride were performed in Earth Sciences Department of Turin University using ion chromatography. Nitrate isotopes were determined by IRMS at ISO4 s.n.c, Italy. Samples were prepared and purified according to the method described by Silva et al. (2000). Uncertainties are ± 0.5‰ for δ15NNO3 and ± 1‰ for δ18ONO3. Boron concentration was determined by ICP-OES (APAT-IRSA 2003) at Nuovi Servizi Ambientali s.r.l., Italy. Boron isotopes (expressed as δ11B‰ vs NBS951) were determined by MC-ICP-MS at ALS Scandinavia AB, Sweden, with an uncertainty of ± 0.4 to ± 1‰.

Results and discussion

Nitrate concentrations in groundwater

The study area displays different degrees of groundwater nitrate contamination (Table 1). The nitrate distribution map of October 2014 is reported in Fig. 3.

In the shallow aquifer the nitrate concentration ranges between 5.3 mg/L and 177.5 mg/L. About 26% of monitored points have nitrate concentration between less than 25 mg/L, 50% between 25 and 50 mg/L, 18% between 50 and 100 mg/L and 6% higher than 100 mg/L up to 177.5 mg/L (Fig. 3). The most polluted areas are located in Poirino Plateau and in Cuneo plain, near Fossano and Racconigi towns. Locally the Turin plain, south to Turin, shows concentration superior than 50 mg/L, near Pinerolo and Carmagnola towns. Near the Po river, the nitrate concentrations are generally low.

These results are consistent with previous studies (Lasagna et al. 2005, 2013, 2015, 2016b; Debernardi et al. 2008).

In deep aquifers all water samples have nitrate level inferior than 50 mg/L, ranging between 5 and 40 mg/L (Fig. 3).

The most polluted shallow groundwater (nitrate concentrations superior than 100 mg/L) is located in NVZs designated in the first phase in 2002 (DPGR 9/R 2002). Actually these areas were selected because already contaminated according to the previous monitoring programs. The sampling points with nitrate levels between 50 and 100 mg/L are generally comprised in the NVZs, with few exceptions.

In the study area the nitrate concentration strictly depends on the well/piezometer depth. In Fig. 4 the nitrate levels were correlated with the sampling point depth and a clear relation was highlighted. More specifically, high NO3 concentrations (> 50 mg/L) are always located in superficial wells, with depth inferior than 30 m. In particular, the highest NO3 levels (> 100 mg/L) are present only in wells with depth lower than 20 m; these points are positioned in the Poirino Plateau.
Fig. 4

Nitrate concentration vs well/piezometer depth both in shallow aquifer and in deep aquifers in Turin-Cuneo plain

As well as at local scale, at regional scale a clear relationship between nitrate concentration and sampling points depth in Piedmont plain was observed by Debernardi et al. (2008) and Lasagna and De Luca (2016).

The nitrite concentrations are inferior than detection limit (0.05 mg/L) in almost all samples, except for an agricultural well (020 in Table 1) and a monitoring well (029 in Table 1) of the Cuneo plain shallow aquifer.

The EC ranges from 222 μS/cm and 1355 μS/cm in the shallow aquifer, and from 407 μS/cm and 608 μS/cm in the deep aquifer. Chloride varies between 3.7 mg/L and 101.2 mg/L in the shallow aquifer and between 5.7 mg/L and 46.2 mg/L in the deep aquifer.

Where sewage is the dominant form of contamination, a close correlation between nitrate and chloride concentrations in groundwater may be observed (Barrett and Howard 2002; Morris et al. 2003). A further cause of nitrate-chloride correlation can be attributed to agricultural activities (Saffigna and Keeney 1977; Kohn et al. 2016), because of the use of synthetic fertilizer, primarily KCl (Kelly et al. 2012). Thus nitrate and chloride were plotted in diagrams of Fig. 4b, showing a moderate correlation between the parameters (R2 = 0.6905, p value = 1.176e-09). However, the correlation analyze is strongly influenced by samples from Poirino Plateau; indeed, performing correlation analyses for the subsets of the data, it was possible to find a low correlation between nitrate and chloride in Turin plain and Cuneo plain, while the correlation in very strong in the Poirino Plateau, with a R2 of 0.9379 (p value = 0.001476).

Nitrate and chloride are often used as indicators of contamination in groundwater. However, they have a different behavior in groundwater: while chloride is a conservative tracer, which typically does not undergo biological transformations or sorb to soil, nitrate can be attenuated in groundwater by denitrification processes in the presence of appropriate bacteria and redox conditions (Korom 1992; Rivett et al. 2008). Thus the presence of this process in Turin-Cuneo groundwater could have reduced the degree of correlation between the parameters in bi-plot of Fig. 5b.
Fig. 5

Plot of nitrate versus electrical conductivity (a) and nitrate versus chloride concentrations (b) in the shallow aquifer of Turin-Cuneo plain

Also the bi-plot nitrate-EC highlights a moderate degree of correlation (R2 = 0.5091, p value = 2.169e-06) (Fig. 5a). Similarly, to the scatter diagram nitrate-chloride, the correlation analyze between nitrate and EC is strongly influenced by samples from Poirino Plateau; particularly there is no correlation between the sample points in Turin and Cuneo plains, while a strong correlation was obserbed in the Poirino Plateau, with a R2 of 0.8554 (p value = 0.008252).

Nitrate isotopes for the identification of sources and fate

The δ15NNO3 values of Turin-Cuneo plain show high variability, especially in the shallow aquifer (Table 1): the values range between 4.00‰ and 20.66‰ in the shallow aquifer and between 6.05‰ and 9.84‰ in deep aquifers.

The δ18ONO3 values range between 4.2‰ and 14.9‰ in the shallow aquifer and between 5.8‰ and 23.6‰ in deep aquifers. More specifically, in the shallow aquifer, better characterized due to the larger number of samples, the δ15NNO3 values of the Cuneo Plain samples vary between 4.00 ‰ and 15.61‰; the δ18O values vary between 4.2‰ and 10.0‰. In the Poirino Plateau, the δ15N NO3 values cover a broad spectrum, ranging between 7.55‰ and 20.66 ‰. The δ18O values vary between 10.0‰ and 13.6‰. Finally, in the Turin Plain, the δ15NNO3 values vary between 5.47‰ and 13.81‰ and the δ18O values between 4.5‰ and 14.9‰.

In order to define the main sources of nitrate in groundwater, isotope data of nitrate were plotted in Fig. 6. The correlation diagram shows that nitrate contamination in the Turin-Cuneo plain originates from: synthetic fertilizers (defined as A in Fig. 6); mixtures of synthetic and organic sources, slightly denitrified (B in Fig. 6); manure or septic tank effluents (C in Fig. 6); soil organic matter (D in Fig. 6).
Fig. 6

Correlation diagram between the δ15NNO3 and δ18ONO3 of dissolved nitrates (after Clark and Fritz 1997). See text for capital letters explanation

However, regarding the samples defined as D in Fig. 6, they show nitrate levels higher than 35 mg/L, and consequently it is more probable to hypothesize their source as a mixture of synthetic and organic sources.

Two samples have nitrate deriving from synthetic fertilizers (A in Fig. 6) and belong respectively to the shallow and deep aquifer. The δ15NNO3 of both of them is higher than +6‰, while generally the δ15NNO3 of synthetic fertilizers is inferior than +4‰ (Clark and Fritz 1997). In general, a total range of −8 to +7‰ was observed by different Authors (Hübner 1986; Macko and Ostrom 1994; Vitoria et al. 2004). These high values can be attributed to nitrate fertilizers that often have slightly higher δ15NNO3 values than NH4+ fertilizers (Kendall et al. 2007).

Both in shallow and in deep aquifer nitrate derives in most of samples from mixtures of synthetic and organic sources, probably with a light degree of denitrification (B in Fig. 6). Sacchi et al. (2013) hypothesized that the same isotopic composition, in water sampled in the central Po Plain, could also derive from mixed sources in disequilibrium with δ18OH2O.

While nitrate from mixed sources (B and D in Fig. 6) mainly belongs to Cuneo and Turin plains, nitrate deriving from manure or sewage (C in Fig. 6) was only sampled in Cuneo plain.

In addition to nitrate sources identification, the correlation analysis δ15NNO3 vs δ18ONO3 was also used to evaluate the denitrification and nitrification processes.

Aravena and Robertson (1998) demonstrated a concomitant enrichment in δ15NNO3 and δ18ONO3 during denitrification. In addition, nitrification produces a positive correlation between the nitrate concentrations and δ 15N NO3 (Choi et al. 2003), due to isotopic fractionation. Indeed, since 14N reacts faster than 15N during nitrification, incomplete nitrification results in 15N-depleted nitrate (Mariotti et al. 1981). Therefore, preferential inflows into groundwater of nitrate produced through incomplete nitrification result in low nitrate concentrations and δ15NNO3. With the proceeding of nitrification to completion, the concentration of nitrate and δ15NNO3 increase gradually. The slope of regression line would be different depending on the initial δ15NNO3 of original source-N (Choi et al. 2003).

A significant denitrification, highlighted by a progressive enrichment in δ15NNO3 and δ18ONO3, was distinguished in the shallow aquifer, especially in the Poirino Plateau (E in Fig. 6a). This result was also highlighted in previous studies analyzing denitrification in the Turin-Cuneo plain (Lasagna et al. 2005, 2016b; Debernardi et al. 2008). The presence of deposits with low permeability in the Poirino Plateau, indeed, can enhance denitrification, as also reported in Sànchez-Pérez et al. (2003); moreover this aquifer has a low oxygenation, also highlighted by high iron and manganese concentrations (Debernardi et al. 2005).

The groundwater samples from the Turin and Cuneo plains generally show a slight denitrification (B in Fig. 6a) or do not show an isotopic trend correlated to a denitrification process.

Furthermore, the positive correlation, especially in the shallow aquifer, between the nitrate concentrations and δ15NNO3 in Turin and Cuneo plains (Fig. 6) supports the presence of a nitrification process in aquifer. This behavior is evident in shallow aquifer for the sampling points of Turin and Cuneo plains (Fig. 7) and it is consistent with the results of diagram of Fig. 6, in which one sample shows an isotopic composition corresponding to synthetic sources, partially nitrified (A in Fig. 6a). A nitrification process is not evident for Poirino Plateau analyzing nitrate vs δ15NNO3 correlation (Fig. 7), due to the scarcity of sampling points.
Fig. 7

δ15NNO3 vs nitrate concentration in the shallow and deep aquifers of Turin-Cuneo plain

Boron concentration in groundwater and boron isotopes

Boron isotopes were used, associated with δ15NNO3, to discriminate further among the main anthropogenic sources of pollution, because fertilizers, manure and wastewater may have characteristic isotopic compositions.

The boron concentration, analyzed in 34 shallow aquifer and 8 deep aquifers samples, is generally inferior than 15 μg/L (detection limit), with the exception of four samples of shallow aquifer in which the concentration ranges between 20.2 and 86.1 μg/L. Two samples belong to Poirino Plateau and the other two to the Turin Plain. These samples were also analyzed for δ11B, showing an isotopic composition from 8.37‰ to 18.05‰ (Table 1). These compositions are consistent with the most frequent range for boron in Italian groundwater, that is comprised between − 10‰ and +25‰ (Pennisi et al. 2013).

The boron concentrations do not correlate with nitrate levels; moreover, the sampling points with the highest nitrate concentrations always show boron content inferior than detection limits.

The δ11B were then plotted with δ15NNO3 in the diagram of Fig. 8. The δ11B values, used for compositional fields, are derived from previous studies (Komor 1997; Widory et al. 2004, 2005; Tirez et al. 2010; Pennisi et al. 2013). More specifically, low δ11B values (about from − 6‰ to + 13‰) are generally related to contamination from sewage and/or synthetic fertilizers, whereas manure (cattle, hog, poultry) has an isotopic signature more positive (about from + 7‰ to + 42‰).
Fig. 8

δ15NNO3 versus δ11B in the shallow aquifers of Turin plain and Poirino Plateau. The number represents the name of the water sample (see Table 1)

According to the obtained results (Table 1), the hypothesized mixed origin of nitrates (sewage and manure) was confirmed. More specifically in Poirino Plateau, nitrate is related to hog and poultry manure. Samples 014 and 017 have nitrate levels equal to 27 and 29 mg/L and δ11B equal to 18.05‰ and 16.77‰ respectively; they were both sampled in agricultural wells with a depth inferior than 20 m. In the Turin plain, the sample 42 was obtained from a 20-m deep monitoring well, located in a rural area. The nitrate concentration is relatively high (41.1 mg/L) and δ11B is 12.99‰. According to Fig. 8, the source of nitrate is cattle manure. The point 41 was also collected in a monitoring well, 42.2 m deep, located in the center of Turin city. The nitrate concentration is 35.8 mg/L and δ11B is 8.37‰. The δ15NNO3 versus δ11B correlation highlights an origin related to sewage. This confirms the presence of leaking sewage networks under big cities.

The origin of the highest nitrate concentrations was not defined using this method because of the low concentration on B in groundwater.

The results of this study are consistent with studies made in similar contexts, e.g., Sacchi et al. (2013) conducted a B isotopic investigation in highly impacted areas of the alpine foothills of Lombardy (NW Italy). In their study, boron concentrations ranged from 20 to 540 μg l−1 and δ 11B values from − 1.4‰ to + 26.2‰ and these results, correlated with isotopic signature of nitrate, confirmed that both sewage and animal manure could contribute to nitrate pollution in the investigated area. However, contrarily to our study, contamination from sewage was observed in areas mainly devoted to agriculture whereas samples from suburban Milan, where sewage was considered the most likely source of contamination, show a δ11B typical for cattle manure. The authors suggest that nitrate in Milan area can be related to contamination that occurred in the past, when the area was still mostly devoted to agriculture.

Conclusions

The use of stable isotope of nitrate and boron can be a very useful tool in the investigation of diffuse contamination. Many nitrate sources can introduce nitrate in groundwater system, and among them sewage, animal manure, chemical fertilizer and natural soil mineralization. To discriminate between them is not simple but extremely useful, especially to implement management actions for groundwater protection.

The study focused on the shallow aquifer of Turin-Cuneo plain. It shows high nitrate levels, having 50% of monitored points nitrate concentration between 25 and 50 mg/L, and 24% superior than 50 mg/L, up to 177 mg/L. The most polluted areas are located in Poirino Plateau and in Cuneo plain, near Fossano and Racconigi towns. The deep aquifer groundwaters are generally more protected, and all water samples have nitrate level inferior than 50 mg/L. The boron concentration is generally inferior than 15 μg/L, except four samples of shallow aquifer in which the concentration ranges between 20 and 86 μg/L.

Stable nitrogen (δ15NNO3) and oxygen (δ18ONO3) isotope data of nitrates were used to define the main sources of nitrates in groundwater. The isotopes investigation indicates that nitrate contamination in the study area originates from different sources, more specifically from mixtures of synthetic and organic sources, and manure and/or septic tank effluents. In effect, the mere use of nitrate stable isotopes is unable to discriminate between manure and sewage effluents. The integrated use of δ15NNO318ONO3 and boron isotope δ11B, instead, can contribute to discriminate further among sewage/septic plumes and cattle, poultry and hog manure sources, because of their different isotopic composition. However, the presence of very low concentration of boron in shallow groundwater permitted to evaluate δ11B only in four water sample, 2 located in Turin plain and 2 in Poirino Plateau, the most contaminated area. The analyses confirm that both the sources are present, precisely hog and poultry manure in Poirino Plateau, cattle manure in Turin plain and sewage under Turin city. This supports the presence of leaking sewage networks under big cities.

The stable isotopes of nitrate were also used to understand the nitrate dynamics in groundwater. The main results are the evaluation of a nitrification process and a subsequent phenomenon of denitrification in aquifer. The identification of denitrification process is very important in hydrogeological studies, because denitrification decreases the nitrate concentrations in groundwater and, therefore, plays a fundamental role in nitrate attenuation. Particularly a significant denitrification process was highlighted in the shallow aquifer of Poirino Plateau that on the other hand is also the most contaminated part of the study area. Moreover, contrary to expectations, a strong denitrification process was not assessed in the deep aquifer.

This study indicates the presence of many source of nitrate contamination, highlighting a significant contribution provided by synthetic fertilizers and manure. High nitrate concentration were observed also in NVZs (Nitrate Vulnerable Zones, according to Nitrate Directive), where the use of manure was significantly restricted (170 kg N ha−1 year−1) according to Nitrate Directive. Currently the limit is 250 kg N ha−1 year−1 after European Commission derogations (European Commission 2011, 2016). The only monitoring point located under a big city, in this case study Turin, underline the possibility of nitrate pollution due to losses from sewage networks. Management plans and remedial actions should consider both agricultural and urban underground infrastructures in order to mitigate the nitrate pollution.

To focus more deeply and comprehensively the problem, future research should focus also on pathogens and emerging contaminants (caffeine, sweeteners, antibiotics…) characterized by a widespread presence in the human diet or habits, to use as environmental tracer. The results of these studies could provide new data to identify sources and pathways of contamination and consequently give useful indications to assist the stakeholders.

Notes

Acknowledgments

This study was supported financially, especially for the nitrate and boron isotopic analyses, by Fondazione Cassa di Risparmio di Torino under the project “Valutazione dell'origine della contaminazione da nitrati nelle acque sotterranee della pianura piemontese.”

References

  1. ADAS (2011) Economics report for NIT18 NVZ action programme impact assessment pp41. https://www.gov.uk/government/uploads/system/uploads/attachment_data/file/82410/20111220nitrates-directive-consult-evid3.pdf
  2. AdBPo (2001). Piano stralcio per l’Assetto Idrogeologico (PAI). D.P.C.M. 24 maggio 2001Google Scholar
  3. Agrawal GD, Lunkad SK, Malkhed T (1999) Diffuse agricultural nitrate pollution of groundwaters in India. Wat Sci Technol 39(3):67–75Google Scholar
  4. Al-Agha MR (1999) Impact of waste water management on groundwater quality in the Gaza Strip, Palestine. In: Chilton (ed) Groundwater in the urban environment: selected city profiles. Balkema, Rotterdam, pp 77–84Google Scholar
  5. APAT-IRSA (2003) Analytical methods for waters (in Italian). Serie APAT Manuali e Linee Guida 29/2003. APAT, RomeGoogle Scholar
  6. Aravena R, Robertson WD (1998) Use of multiple isotope tracers to evaluate denitrification in ground water: study of nitrate from a largeflux septic system plume. Ground Water 36(6):975–982.  https://doi.org/10.1111/j.1745-6584.1998.tb02104.x Google Scholar
  7. Aravena R, Evans ML, Cherry JA (1993) Stable isotopes of oxygen and nitrogen in source identification of nitrate from septic tanks. Ground Water 31(2):180–186.  https://doi.org/10.1111/j.1745-6584.1993.tb01809.x Google Scholar
  8. Baker L (1992) Introduction to nonpoint source pollution in the United States and prospects for wetland use. Ecol Eng 1(1-2):1–26.  https://doi.org/10.1016/0925-8574(92)90023-U Google Scholar
  9. Barbero D, De Luca DA, Forno MG, Lasagna M (2016) Preliminary results on temperature distribution in the quaternary fluvial and outwash deposits of the Piedmont Po Plain (NW Italy): a statistical approach. Rend Online Soc Geol Ital 41(2016):272–275.  https://doi.org/10.3301/ROL.2016.146 Google Scholar
  10. Barrett MH, Howard AG (2002) Urban groundwater and sanitation—developed and developing countries. In: KWF H, Israfilov RG (eds) Current problems of hydrogeology in urban areas. Urban agglomerates and industrial centres. Kluwer Academic, Amsterdam, pp 39–56Google Scholar
  11. Bassett RL, Buszka PM, Davidson GR, Chong-Diaz D (1995) Identification of groundwater solute sources using boron isotopic composition. Environ Sci Technol 29(12):2915–2922.  https://doi.org/10.1021/es00012a005 Google Scholar
  12. Bortolami GC, Maffeo B, Maradei V, Ricci B, Sorzana F (1976) Lineamenti di litologia e geoidrologia del settore piemontese della Pianura Padana. Quaderni dell’Istituto di Ricerca sulle Acque 28(1):3–37 RomaGoogle Scholar
  13. Bove A, Casaccio D, Destefanis E, De Luca DA, Lasagna M, Masciocco L, Ossella L, Tonussi M (2005) Idrogeologia della pianura piemontese, Regione Piemonte. Mariogros Industrie Grafiche S.p.A, Torino (CD-Rom)Google Scholar
  14. Bucci A, Barbero D, Lasagna M, Forno MG, De Luca DA (2017) Shallow groundwater temperature in the Turin area (NW Italy): vertical distribution and anthropogenic effects. Environ Earth Sci 76(5):221.  https://doi.org/10.1007/s12665-017-6546-4 Google Scholar
  15. Canavese PA, De Luca DA, Masciocco L (2004) La rete di monitoraggio delle acque sotterranee delle aree di pianura della Regione Piemonte: quadro idrogeologico. Prismas: il monitoraggio delle acque sotterranee nella Regione Piemonte. Mariogros Industrie Grafiche S.p.A., Torino, 180 ppGoogle Scholar
  16. Capri E, Civita M, Corniello A, Cusimano G, De Maio M, Ducci D, Fait G, Fiorucci A, Hauser S, Pisciotta A, Pranzini G, Trevisan M, Delgado Huertas A, Ferrari F, Frullini R, Nisi B, Offi M, Vaselli O, Vassallo M (2009) Assessment of nitrate contamination risk: the Italian experience. J Geochem Explor 102(2):71–86.  https://doi.org/10.1016/j.gexplo.2009.02.006 Google Scholar
  17. Castagna S, Dino GA, Lasagna M, De Luca DA (2015a) Environmental issues connected to the quarry lakes and chance to reuse fine materials deriving from aggregate treatments. In: Lollino G et al (eds) Engineering geology for society and territory–volume 5, urban geology, sustainable planning and landscape exploitation. Springer International Publishing, Switzerland, pp 71–74.  https://doi.org/10.1007/978–3–319-09048-1_13 Google Scholar
  18. Castagna SED, De Luca DA, Lasagna M (2015b) Eutrophication of Piedmont quarry lakes (north-western Italy): hydrogeological factors, evaluation of trophic levels and management strategies. J Environ Assess Pol Manag 17:4 (21 pages), Imperial College Press.  https://doi.org/10.1142/S1464333215500362 Google Scholar
  19. Choi WJ, Lee SM, Ro HM (2003) Evaluation of contamination sources of groundwater NO3 using nitrogen isotope data: a review. Geosci J 7(1):81–87.  https://doi.org/10.1007/BF02910268 Google Scholar
  20. Chowdary VM, Rao NH, Sarma PBS (2005) Decision support framework for assessment of non-point-source pollution of groundwater in large irrigation projects. Agric Water Manag 75(3):194–225.  https://doi.org/10.1016/j.agwat.2004.12.013 Google Scholar
  21. Clark I, Fritz P (1997) Environmental isotopes in hydrogeology. Lewis Publishers, New York, p 328Google Scholar
  22. Clemente P, Lasagna M, Dino GA, De Luca DA (2015) Comparison of different methods for detecting irrigation canals leakage. In: Lollino G et al (eds) Engineering geology for society and territory–volume 3, River Basins, reservoir sedimentation and water resources. Springer International Publishing, Switzerland, pp 23–26.  https://doi.org/10.1007/978–3–319-09054-2_5 Google Scholar
  23. Comina C, Lasagna M, De Luca DA, Sambuelli L (2014) Geophysical methods to support correct water sampling locations for salt dilution gauging. Hydrol Earth Syst Sci 18(8):3195–3203.  https://doi.org/10.5194/hess-18-3195-2014 Google Scholar
  24. De Luca DA, Lasagna M, Morelli di Popolo e Ticineto A (2007) Installation of a vertical slurry wall around an Italian quarry lake: complications arising and simulation of the effects on groundwater flow. Environ Geol 53(1):177–189.  https://doi.org/10.1007/s00254-006-0632-3 Google Scholar
  25. De Luca DA, Dell’Orto V, Destefanis E, Forno MG, Lasagna M, Masciocco L (2009) Hydrogeological structure of the “fontanili” in Turin plain. Rend Online Soc Geol Ital 6:199–200Google Scholar
  26. De Luca DA, Destefanis E, Forno MG, Lasagna M, Masciocco L (2014) The genesis and the hydrogeological features of the Turin Po Plain fontanili, typical lowland springs in northern Italy. Bull Eng Geol Environ 73:409–427.  https://doi.org/10.1007/s10064-013-0527-y Google Scholar
  27. Debernardi L., De Luca DA, Lasagna M (2005) Il processo di denitrificazione naturale nelle acque sotterranee in Piemonte “natural denitrification in groundwater in the western sector of the Po plain (northern Italy)”. Proceedings of aquifer vulnerability and risk, 2nd international workshop and 4th National Congress on the protection and Management of Groundwater-Reggia di Colorno (PR), Italy, 21–23 September 2005, paper ID 176, 27 ppGoogle Scholar
  28. Debernardi L, De Luca DA, Lasagna M (2008) Correlation between nitrate concentration in groundwater and parameter affecting aquifer intrinsic vulnerability. Environ Geol 55(3):539–558.  https://doi.org/10.1007/s00254-007-1006-1 Google Scholar
  29. DGR 23–13437 (2004) Decreto legislativo 11 maggio 1999 n. 152 art. 44. Adozione del Piano regionale di tutela delle acque (PTA) e proposta al Consiglio Regionale della relativa approvazione. Suppl. al B.U. n. 44 del 4 novembre 2004Google Scholar
  30. DGR. 28–2845 (2006). Modifiche e integrazioni alla D.G.R. 20 settembre 2004 n. 23–13437 (come modificata dalla d.g.r. 17 gennaio 2005, n. 30–14577) di adozione del Piano di tutela delle acque e proposta al Consiglio Regionale della relativa approvazione. B.U. n. 21 del 25 / 05/2006Google Scholar
  31. DPGR 12/R (2007) Designazione di ulteriori zone vulnerabili da nitrati di origine agricola ai sensi del decreto legislativo 3 aprile 2006, n. 152. (Legge regionale 29 dicembre 2000, n. 61). B.U. Regione Piemonte, numero 1–3 gennaio 2008Google Scholar
  32. DPGR 9/R (2002) Regolamento regionale recante: Designazione delle zone vulnerabili da nitrati di origine agricola e relativo programma d’azione. Bollettino Ufficiale Regione Piemonte, 2° Suppl. al numero 43–24 ottobre 2002Google Scholar
  33. European Commission (1991) Directive 91/676/EEC. Council directive of 12 December 1991 concerning the protection of waters against pollution caused by nitrates from agricultural sources. Off J Eur Communities 375:1–8 Available at: http://eur-lex.europa.eu/LexUriServ/LexUriServ.do?uri=OJ:L:1991:375:0001:0008:EN:PDF Google Scholar
  34. European Commission (2011) Commission implementing decision of 3 November 2011 on granting a derogation requested by Italy with regard to the regions of Emilia Romagna, Lombardia, Piemonte and Veneto pursuant to Council Directive 91/676/EEC concerning the protection of waters against pollution caused by nitrates from agricultural sources (2011/721/EU). Off J Eur Union 287:36–41 Available at: http://eur-lex.europa.eu/LexUriServ/LexUriServ.do?uri=OJ:L:2011:287:0036:0041:EN:PDF Google Scholar
  35. European Commission (2016) Commission Implementing Decision (EU) 2016/1040 of 24 June 2016 on granting a derogation requested by Italian Republic with regard to the Regions of Lombardia and Piemonte pursuant to Council Directive 91/676/EEC concerning the protection of waters against pollution caused by nitrates from agricultural sources (notified under document C(2016) 3820)Google Scholar
  36. Harter T, Davis H, Mathews M, Meyer R (2002) Shallow groundwater quality on dairy farms with irrigated forage crops. J Contam Hydrol 55(3-4):287–315.  https://doi.org/10.1016/S0169-7722(01)00189-9 Google Scholar
  37. Hübner H (1986) Isotope effects of nitrogen in the soil and biosphere. In: Fritz P, Fontes JC (eds) Handbook of environmental isotope geochemistry, Vol 2b, the terrestrial environment. Elsevier, Amsterdam, pp 361–425Google Scholar
  38. Kelly WR, Panno SV, Hackley KC (2012) The sources, distribution, and trends of chloride in the waters of Illinois. Illinois State Water Survey, Prarie Research Institute, Champaign, p 67 (Mar, Report No.: B-74)Google Scholar
  39. Kendall C (1998) Tracing nitrogen sources and cycling in catchment. In: Kendall C, McDonnell JJ (eds) Isotope tracers in catchment hydrology. Elsevier, Amsterdam, pp 519–576.  https://doi.org/10.1016/B978-0-444-81546-0.50023-9 Google Scholar
  40. Kendall C, Elliott EM, Wankel SD (2007) Tracing anthropogenic inputs of nitrogen to ecosystems, chapter 12. In: Michener RH, Lajtha K (eds) Stable isotopes in ecology and environmental science, 2nd edition. Blackwell Publishing, Hoboken, pp 375–449.  https://doi.org/10.1002/9780470691854.ch12 Google Scholar
  41. Kohn J, Soto DX, Iwanyshyn M, Olson B, Kalischuk A, Lorenz K, Hendry MJ (2016) Groundwater nitrate and chloride trends in an agriculture intensive area in southern Alberta, Canada. Water Qual Res J 51(1):47–59.  https://doi.org/10.2166/wqrjc.2015.132 Google Scholar
  42. Komor SC (1997) Boron contents and isotopic compositions of hog manure, selected fertilizers, and water in Minnesota. J Environ Qual 26(5):1212–1222.  https://doi.org/10.2134/jeq1997.00472425002600050004x Google Scholar
  43. Korom SF (1992) Natural denitrification in the saturated zone: a review. Water Resour Res 28(6):1657–1668.  https://doi.org/10.1029/92WR00252 Google Scholar
  44. Lasagna M, De Luca DA (2016) The use of multilevel sampling techniques for determining shallow aquifer nitrate profiles. Environ Sci Pollut Res 23:20431–20448.  https://doi.org/10.1007/s11356-016-7264-2 Google Scholar
  45. Lasagna M, De Luca DA, Sacchi E, Bonetto S (2005) Studio dell’origine dei nitrati nelle acque sotterranee piemontesi mediante gli isotopi dell’azoto. Giornale di Geologia Applicata 2:137–143Google Scholar
  46. Lasagna M, De Luca DA, Debernardi L, Clemente P (2013) Effect of the dilution process on the attenuation of contaminants in aquifers. Environ Earth Sci 70(6):2767–2784.  https://doi.org/10.1007/s12665-013-2336-9 Google Scholar
  47. Lasagna M, Caviglia C, De Luca DA (2014) Simulation modeling for groundwater safety in an overexploitation situation: the Maggiore Valley context (Piedmont, Italy). Bull Eng Geol Environ 73:341–355.  https://doi.org/10.1007/s10064-013-0500-9 Google Scholar
  48. Lasagna M, Franchino E, De Luca DA (2015) Areal and vertical distribution of nitrate concentration in Piedmont plain aquifers (North-western Italy). In: Lollino G et al (eds) Engineering geology for society and territory–volume 3, River Basins, reservoir sedimentation and water resources. Springer International Publishing, Switzerland, pp 389–392.  https://doi.org/10.1007/978–3–319-09054-2_81 Google Scholar
  49. Lasagna M, De Luca DA, Franchino E (2016a) Nitrate contamination of groundwater in the western Po Plain (Italy): the effects of groundwater and surface water interactions. Environ Earth Sci 75(3):240.  https://doi.org/10.1007/s12665-015-5039-6. Google Scholar
  50. Lasagna M, De Luca DA, Franchino E (2016b) The role of physical and biological processes in aquifers and their importance on groundwater vulnerability to nitrate pollution. Environ Earth Sci 75(11):961.  https://doi.org/10.1007/s12665-016-5768-1 Google Scholar
  51. Liao L, Green CT, Bekins BA, Böhlke JK (2012) Factors controlling nitrate fluxes in groundwater in agricultural areas. Water Resour Res 48:W00L09.  https://doi.org/10.1029/2011WR011008 Google Scholar
  52. Macko SA, Ostrom NE (1994) Pollution studies using nitrogen isotopes. In: Lajtha K, Michener RM (eds) Stable isotopes in ecology and environmental science. Blackwell Scientific Publishers, Oxford, pp 45–62Google Scholar
  53. MacQuarrie KTB, Sudicky EA, Robertson WD (2001) Numerical simulation of a fine-grained denitrification layer for removing septic system nitrate from shallow groundwater. J Contam Hydrol 52:29–55Google Scholar
  54. Mariotti A, Germon JC, Hubert P, Kaiser P, Letolle R, Tardieux A, Tardieux P (1981) Experimental determination of nitrogen kinetic isotope fractionation: some principles; illustration for the denitrification and nitrification processes. Plant Soil 62:423–430Google Scholar
  55. Ministry for Environment, Land and Sea, Ministry for Agriculture, Food and Forestry Policies, Regions of Piedmont, Lombardy, Veneto, Emilia-Romagna and Friuli Venezia Giulia (2010). Request from Italy for a derogation under paragraph 2(b) of Annex III to directive 91/676/EEC from the limit of 170 kilograms of nitrogen per hectare per year from livestock manure. http://www.regione.piemonte.it/agri/politiche_agricole/dirett_nitrati/dwd/07102011/Documento_tecnico_scientifico.pdf
  56. Morris BL, Lawrence ARL, Chilton PJC, Adams B, Calow RC, Klinck BA (2003) Groundwater and its susceptibility to degradation: a global assessment of the problem and options for management. Early warning and assessment report series, RS. 03-3. United Nations Environment Programme, NairobiGoogle Scholar
  57. Nolan B, Stoner J (2000) Nutrients in groundwaters of the conterminous United States, 1992–1995. (2000). USGS staff—published research. Paper 59. http://digitalcommons.unl.edu/usgsstaffpub/59
  58. Pennisi M, Adorni-Braccesi A, Andreani D, Gori L, Sciuto PF, Gonfiantini R (2013) ISOBORDAT: an online database on boron isotopes. Proc. Internat. Symp. Isotopes in Hydrology, Marine Ecosystems, and Climate Change Studies, Oceanographic Museum, Principality of Monaco, 27 March–1 April 2011. IAEACN-186-061Google Scholar
  59. Perotti L, Lasagna M, Clemente P, Dino GA, De Luca DA (2015) Remote sensing and hydrogeological methodologies for irrigation canal water losses detection: the Naviglio di Bra test site (NW-Italy). In: Lollino G et al (eds) Engineering geology for society and territory–volume 3, River Basins, reservoir sedimentation and water resources. Springer International Publishing, Switzerland, pp 19–22.  https://doi.org/10.1007/978–3–319-09054-2_4 Google Scholar
  60. Petitta M, Fracchiolla D, Aravena R, Barbieri M (2009) Application of isotopic and geochemical tools for the evaluation of nitrogen cycling in an agricultural basin, the Fucino Plain, Central Italy. J Hydrol 372(1-4):124–135.  https://doi.org/10.1016/j.jhydrol.2009.04.009 Google Scholar
  61. Postma D, Boesen C, Kristiansen H, Larsen F (1991) Nitrate reduction in an unconfined aquifer: water chemistry, reduction processes, and geochemical modeling. Water Resour Res 27(8):2027–2045.  https://doi.org/10.1029/91WR00989 Google Scholar
  62. Re V, Sacchi E (2017) Tackling the salinity-pollution nexus in coastal aquifers from arid regions using nitrate and boron isotopes. Environ Sci Pollut Res 24(15):13247–13261.  https://doi.org/10.1007/s11356-017-8384-z Google Scholar
  63. Regione Piemonte (2008) Carta dell’uso del suolo 1:500000. Available at: http://www.regione.piemonte.it/territorio/dwd/pianifica/tavoloInterregionale/usoSuolo.pdf. Accessed 29 July 2015
  64. Rivett MO, Buss SR, Morgan P, Smith JWN, Bemment CD (2008) Nitrate attenuation in groundwater: a review of biogeochemical controlling processes. Water Res 42(16):4215–4232.  https://doi.org/10.1016/j.watres.2008.07.020 Google Scholar
  65. Sacchi E, Pilla G, Gerbert-Gaillard L, Zuppi GM (2007) A regional survey on nitrate contamination of the Po valley alluvial aquifer (northern Italy). Int. Symp. On advances in isotope hydrology and its role in sustainable water resources management, IAEA, Vienna 21-25 May 2007, IAEA-CN-151/34, vol., 2, 471-478Google Scholar
  66. Sacchi E, Acutis M, Bartoli M, Brenna S, Delconte CA, Laini A, Pennisi M (2013) Origin and fate of nitrates in groundwater from the central Po plain: insights from isotopic investigations. Appl Geochem 34:164–180.  https://doi.org/10.1016/j.apgeochem.2013.03.008 Google Scholar
  67. Sacco D, Zavattaro L, Grignani C (2006) Regional-scale predictions of agricultural n losses in an area with a high livestock density. Ital J Agron 4:689–703Google Scholar
  68. Saffigna PG, Keeney DR (1977) Nitrate and chloride in ground water under irrigated agriculture in Central Wisconsin. Groundwater 15(2):170–177.  https://doi.org/10.1111/j.1745-6584.1977.tb03162.x Google Scholar
  69. Sànchez-Pérez JM, Bouey C, Sauvage S, Teissier S, Antiguedad I, Vervier P (2003) A standardised method for measuring in situ denitrification in shallow aquifers: numerical validation and measurements in riparian wetlands. Hydrol Earth Syst Sci 7(1):87–96.  https://doi.org/10.5194/hess-7-87-2003 Google Scholar
  70. Seiler RL (2005) Combined use of 15N and 18O of nitrate and 11B to evaluate nitrate contamination in groundwater. Appl Geochem 20(9):1626–1636.  https://doi.org/10.1016/j.apgeochem.2005.04.007 Google Scholar
  71. Silva SR, Kendall C, Wilkison DH, Ziegler AC, Chang CCY, Avanzino RJ (2000) A new method for collection of nitrate from fresh water and the analysis of nitrogen and oxygen isotope ratios. J Hydrol 228(1-2):22–36.  https://doi.org/10.1016/S0022-1694(99)00205-X Google Scholar
  72. Tirez K, Brusten W, Widory D, Petelet E, Bregnot A, Xue D, Boeckx P, Bronders J (2010). Boron isotope ratio (d11B) measurements in Water Framework Directive monitoring programs: comparison between double focusing sector field ICP and thermal ionization mass spectrometry. J Anal At Spectrom 25:964–974Google Scholar
  73. Vengosh A (1998) Boron isotopes and groundwater pollution. Water Environ News 3:15–16Google Scholar
  74. Vengosh A, Heumann KG, Juraske S, Kasher R (1994) Boron isotope application for tracing sources of contamination in groundwater. Environ Sci Technol 28(11):1968–1974.  https://doi.org/10.1021/es00060a030 Google Scholar
  75. Vitoria L, Otero N, Soler A, Canals A (2004) Fertilizer characterization: isotopic data (N, S, O, C, and Sr). Environ Sci Technol 38(12):3254–3262Google Scholar
  76. Widory D, Kloppmann W, Chery L, Bonnin J, Rochdi H, Guinamant J (2004) Nitrate in groundwater: an isotopic multi-tracer approach. J Contam Hydrol 72:165–188Google Scholar
  77. Widory D, Petelet-Giraud E, Negrel P, Lafdoich B (2005) Tracking the sources of nitrates in groundwater using coupled nitrogen and boron isotopes: a synthesis. Environ Sci Technol 39(2):539–548.  https://doi.org/10.1021/es0493897 Google Scholar

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© Springer-Verlag GmbH Germany, part of Springer Nature 2017

Authors and Affiliations

  1. 1.Earth Sciences DepartmentTurin UniversityTurinItaly

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