Modern conventional and intensive agricultural prac­tices have an impact on the environment, causing a re­duc­tion of biodiversity, an increase in soil erosion and pollution of underground and surface waters ( Palma et al. , 2007 ; FAO, 2013 ). Excessive application of nitrates in agriculture (inorganic N fertilisers and manures) can leach into groundwater and subsequently negatively impact the environment and pose human health problems ( Leach et al. , 2004 ; Giles, 2005 ; Gutierrez et al. , 2016 ; Pacheco & Fernandes, 2016 ).

The European Commission Council Directive 91/676/EEC (Nitrates Directive), integrated in Directive 2000/60/EC of the European Parliament and of the Council (Framework for the Community action in the field of water policy), indicates the need to introduce measures to protect waters from pollution caused by nitrates deriving from agricultural sources ( Cerro et al. , 2014 ; Lawniczak et al. , 2016 ) and identifies Nitrate Vulnerable Zones (NVZs) as areas of land or water at risk of pollution ( Arauzo & Valladolid, 2013 ), where fertiliser application procedures should be reviewed. In order to achieve the objective of these Directives, Member States are required to implement all necessary measures to prevent or limit inputs of any hazardous substances into groundwater.

In compliance with these directives, the Spanish Royal Decree 261/1996 provides a list of "Codes of Good Agricultural Practices" for farmers, to be implemented generically on a voluntary basis and "Action Programmes" to be implemented specifically within the NVZs on a compulsory basis ( Velthof et al. , 2014 ). These measures limit the periods when N fertilisers can be applied, in order to prevent nutrient leaching and also limit the maximum amount of livestock manure that can be applied, corresponding to 170 kg N ha ^-1 year ^-1 ( Bayo et al. , 2012 ).

Spain, Italy and the south of France are among the major pork producing countries of the EU Mediterranean area ( EC, 2013 ), and they present similar issues with N pollution of ground water due to the lack of a correct management of manure originating from the pork industry ( INE, 2009 ). Although the majority of the EU state members have decreased the number of reared pigs, this has not occurred in the Mediterranean countries.

Manure production in Spain, generated by stabled animals, is 76 million tonnes per year ( Bigeriego, 2001 ). This large volume of liquid manure, especially pig manure, applied to areas of intensive agriculture is creating a significant nitrate pollution of groundwater problem ( Prapaspongsa et al. , 2010 ).

The number of farms and swine bred for the Spanish pork industry has increased over the last decades according to MAPAMA (2016) , reaching a total of 29M swine in 2016 and makes Spain the second largest pork producing country in the EU. In Spain, Catalonia represents the top-ranking region for swine with 7.6 million animals (26% of all Spanish stock). Catalonia produces 43% of all Spanish pork meat and represents 64% of Spanish swine exports (almost 1 million tons). The annual export revenue of pork meat in Catalonia is approx. 2 billion € ( DARPA, 2017 ). Although the number of intensive farms has decreased by 16% over the last 10 years, it is still the most common method of pig rearing and fattening in Spain, with a total number of 68,980 intensive farms versus 14,213 extensive farms ( MAPAMA, 2016 ). Regions such as Andalucía, Ara­gón, Castilla León, Extremadura and Catalonia present a similar number of intensive farms. Conversely, of these regions, Extremadura and Andalusia are the only areas where a significant number of extensive farms are also present (12,084) ( MAPAMA, 2016 ). The pork industry in Catalonia can be defined as intensive, totalling 6,045 farms in 2016, with a high concentration of stabled animals, equal to over 460 animals per farm ( Food & Water Europe, 2017 ).

Pig manure used as a fertiliser is not the only cause of N contamination of ground water. Intensive agriculture, using mineral fertilisers, also produces contamination; however, a clear relationship between pork producing areas, in Spain and Catalonia, and the contamination of ground water exists. As a matter of fact, all the pork production areas in Spain (with the relevant exception of Galicia) report N contamination of groundwater and a part of their territory has been designated as nitrate vulnerable zones ( Food & Water Europe, 2017 ).

Studies on the quality of ground and surface waters of Catalonia, indicate that 38% is contaminated by nitrates and that 34% of the surface has been designated as nitrate vulnerable zones; over the 2000-2010 period, more than 42M € was spent, in order to reduce or to prevent water pollution ( CWA, 2016 ). The nitrate problem in Catalonia has led to the preparation of a new regional decree regarding the management of soil fertilization and manure, which is due to become effective within 2018 and various measures have also been implemented to reduce N leaching, including: catch crops, agroforestry systems, wetlands and biogas plants ( Børgesen & Vinther, 2012 ). Currently, biogas plants are not operative due to their low feasibility and their practicality is subject to an unstable policy of support in Spain ( Capodaglio et al. , 2016 ).

Furthermore, over the last fifty years, Spain has lost 60% of its buffer areas (ditches, rows of trees, hedges, wetlands), which, combined with the evolution of agricultural practices, has led to a consequent increase in erosion, runoff and drainage of nutrients ( Paris et al. , 2002 ).

The objectives of N management are both agrono­mical and environmental and they obviously have eco­nomic consequences. Many authors ( Grignani et al. , 2009 ; Oenema et al. , 2009 ; Cardenas et al. , 2011 ; Oenema, 2012 ; Van der Straeten et al. , 2012 ; Zavattaro et al. , 2012 ) agree that N management measures require additional activities ( i.e. adjustment and reduction of the rate of N fertiliser applied, adjustment of stocking rate and N surplus in animal diets, reduction of the length of the grazing season, adoption of catch crops and irrigation control).

Agroforestry, which is the concurrent cultivation of trees and agricultural crops in the same area, can play a key role in reducing the environmental impact of conventional modern agriculture ( Eichhorn et al. , 2006 ; Rigueiro-Rodríguez et al. , 2008 ; Dougherty et al. , 2009 ; Zamora et al. , 2009 ; Christen & Dalgaard, 2013 ; Smith et al. , 2013 ), however farmers’ willingness to adopt these practices depends on economic and management factors ( Buckley et al. , 2012 ). Recent estimates indicate that the cultivated territory, classified as agroforestry in the EU 27, is approximately 15 million hectares ( den Herder et al. , 2017 ) and Spain, with 5.6 million hectares, has the largest area. Thanks to an integrated land-use system, agroforestry can pro­vide a wide variety of ecosystem services ( Torralba et al. , 2016 ) and social and economic benefits ( Graves et al. , 2007 ). Regarding ecosystem services, agroforestry can contribute to climate change mitigation ( Palma et al. , 2015 ), by reducing CO ₂ emissions produced by agricultural activities, by improving C fixation ( Paolotti et al. , 2016 ), by protecting the soil (both physically and chemically), by improving water quality and by protecting biodiversity ( Liagre & Dupraz, 2008 ).

Agroforestry systems could provide a partial solu­tion for N contamination ( Mosquera-Losada et al. , 2016 ); Palma (2006) estimates that an agroforestry system can reduce up to 28% of N leaching. According to Briggs (2012) , N leaching in silvoarable systems is reduced by 50% compared to monocultures, because the tree roots absorb the excess N not utilised by the arable crop. A more optimistic study ( Dupraz et al. , 2005 ), indicates an approximate 65% reduction in leaching after 60 years and can be considered as a theoretical maximum impact of agroforestry on N leaching.

In accordance with several authors ( Nair et al. , 2007 ; Palma et al. , 2007 ; Christensen et al. , 2013 ), the ability of agroforestry systems to reduce the loss of N can be conservatively estimated from 10% to 50%. Two types of agroforestry practices ( Mosquera-Losada et al. , 2009 ) are particularly efficient in this regard: alley-cropping (AC) and riparian buffer (RB) strips.

Alley cropping consists in combining an arable crop with trees planted in rows ( Burgess et al. , 2005 ; Cardinael et al. , 2015 ), whereas RBs consist of groups of high density trees, or shrubs, established between the arable area and a body of water ( Borin & Bigon, 2002 ), simulating the buffer effect of riparian forests. In both cases, the tree component normally consists of a species providing an income that is either frequent ( i.e. biomass) or significant ( i.e. valuable timber) and can be considered as an activity able to stimulate rural development, by promoting the management of marginal areas through sustainable revenues ( Rois-Díaz et al. , 2017 ).

In fact, as demonstrated by Graves et al. (2017a) and Rois-Diaz et al. (2017) , the obstacles to the diffusion of agroforestry are mainly linked to the lack of profitability. In addition, farmers who choose to use their lands for agroforestry, consider profitability as the main prerequisite, followed by diversification, environmental issues and landscape benefits. An in­crease in productivity and economic outcomes has already been reported ( Herzog, 1998 ; Colomb, 2009 ; Sereke, 2012 ), but the latter strictly depend on the legislative framework; for instance, in the last Common Agricultural Policy (CAP) reform, the threshold of maximum tree density to maintain the land eligible for public funding increased from 50 to 100 trees ha ^-1 . This is still a rather low value for new agroforestry systems, if age is not considered, and has yet to be adopted in many countries, including Spain ( Mosquera-Losada et al. , 2015 ). Other potential lines of subsidy in the most recent CAP were those related to greening practices (Regulation 1307/2013), including, among others, crop diversification and ecological focus areas.

In the light of the above framework, the aim of this paper was to evaluate the profitability of agroforestry practices located in NVZs in a Mediterranean area, in Catalonia, NE Spain, evaluating the competitiveness of these practices in comparison to conventional crops. The economic performance was assessed by exami­ning the various experimental plots, which differed in terms of agricultural cultivations (typical of the examined area and representative for the Spa­nish Mediterranean area), category of agroforestry, tree components, tree density and the presence or absence of irrigation. In addition, the potential of silvoarable systems was also evaluated in relation to the availability of subsidies for the studied area.

The economic comparison of silvoarable systems and arable scenarios refers to EAV for long-term period rotations and to profit for annual crops.

As stated by Sereke et al. (2016) , there is a lack of studies focussing on the socio-economic aspects regarding profitability of agroforestry farming. The present study assessed the profitability of agroforestry at plot scale, as an income opportunity for farmers located in rural areas with environmental problems.

In fact, one of the main drivers of abandonment and decline of rural areas is related to agriculture profitability ( Breustedt & Glauben, 2007 ; Renwick et al. , 2013 ), hence, the first motivation for farmers to start agroforestry practices is of economic nature. Therefore, agroforestry could represent an interesting opportunity, useful to stimulate rural areas by providing additional revenues and employment opportunities ( Rancane et al. , 2014 ; Rois-Díaz et al. , 2017 ). With reference to the Mediterranean area, the main drivers for practicing agroforestry are: diversification of products, improving the environment, life quality and subsidies ( Mosquera-Losada et al. , 2017 ; Rois-Díaz et al. , 2017 ).

Many authors ( i.e. , Graves et al. , 2017b ) state the negative perceptions of silvoarable systems due to intercrop production and timber quality. In our simulation, the results in alley-cropping scenarios confirmed the sharp decline of the agricultural crop component after 20 years, and these results would only be offset by the profits from the tree component at the end of the rotation. Furthermore, agroforestry only appeared to be applicable advantageously in AC with barley, while the scenarios with wheat indicated negative results. Moreover, the results show that to achieve high profitability of the tree component, the objective is not high productivity but rather high-quality timber, such as veneer.

Other authors point out that, in a European environment, agroforestry systems become econo­mically interesting by reducing the number of plants per hectare and with crops lasting between 30 and 60 years ( Sereke et al. , 2015 ), in other environments the agroforestry system is economically viable over 20 years ( Martinelli et al. , 2019 ). It is therefore necessary to highlight that the economic results of the agroforestry system examined in this study are strongly influenced by the ecological and regulatory function for which this type of activity is undertaken. The sensitivity analyses indicated that both AC scenarios only become a competitive option, compared to traditional crops, when agricultural production is reduced by 40-45% and with higher timber prices at 800-850 € m ^-3 . These conditions present a pessimistic scenario for the agricultural component and optimistic for the tree component, however when compared to the experience gained in other contexts it could be considered realistic ( Valdivia et al. , 2012 ; Rancane et al. , 2014 ).

Graves et al. (2017b) highlight that farmers could possibly perceive the advantage of systems located in a Mediterranean context, where the annual crops are obtained from the trees ( i.e. firewood). In this study, the economic results for the RB scenarios indicated that high-density crops involved high production costs that were not offset by the biomass yield. These results are in line with those obtained by Faasch & Patenaude (2012) , who also achieved similar results in Central Europe, reporting high production costs not offset by revenues, and highlighting the importance of subsidies to encourage this type of agricultural activity. The sensitivity analyses conducted for the two RB scenarios also expressed the difficulty of obtaining positive economic results.

Farmers consider that the major benefits of silvoarable systems would be environmental ( Pannell, 1999 ) and previous studies have demonstrated the environmental and social benefits of agroforestry ( Palma et al. , 2007 ). Both silvoarable agroforestry systems studied, in addition to being a business opportunity, provide habitats and refuge zones for birds and other animals, and ensure ecological functions. However, RBs provide important environmental functions, and the loss of profits may be comparable to the cost of deployment of other systems for N pollution reduction ( i.e. reduction at source and treatment of slurry, artificial drainage). This solution could have a positive reception, particularly in rural areas and where water is readily available.

As highlighted, no decrease is expected for the number of stabled animals in the Mediterranean area, and several manure management solutions are needed, moreover farmers with environmental problems are more receptive to the adoption of silvoarable agroforestry ( Graves et al. , 2017b ). Hence, for farmers located in Mediterranean area, the mitigation of groundwater contamination and N leaching could be transformed from an environmental problem to an opportunity.

The main threats to the implementation of the examined practices appear to be high insecurity and dependence on EU subsidies. As also reported by Mosquera-Losada et al. (2017) , in order to make this possible, agroforestry has to be supported by policy and actions, bearing in mind that agroforestry is a land management option that delivers market and non-market goods and services. In both the examined agroforestry categories, it appears that direct payments of subsides for annual crops are fundamental for the success of agriculture in rural contexts. Unlike other six Spanish regions, Catalonia has not activated grants for agroforestry within its Regional Rural Development Programme (sub-measure 8.2) within Pillar II of the CAP. In this regard, Spain operates as a federal state in agricultural management and policies, meaning that Catalonia has the authority to develop its own policy. Thus, support in the form of an annual grant would be justified by the potential of agroforestry systems to reduce N leaching. More generally, the access to forms of subsidies, such as payments for ecosystem services, could be key to encouraging farmers to provide environmental service benefits.

In Pillar I of the CAP, agroforestry is eligible for basic payments, where the density of trees with a crown diameter larger than 4 m is below 100 trees ha ^-1 ( Coello et al. , 2018 ). Moreover, agroforestry practices are eligible to receive payments linked to greening, as they have been activated by Spain as an eligible Ecological Focus Area. Thanks to our simulation, an additional subsidy to that already granted, equal to 122 € ha ^-1 year ^-1 , could be sufficient to make AC economically viable, where the annual crop is wheat, and a subsidy of 137 € ha ^-1 year ^-1 for RBs. A similar picture emerges in other contexts ( Buchholz & Volk, 2013 ), where the profitability of this type of cultivation is strongly affected by incentive programs. Therefore, agroforestry can represent an opportunity in rural areas, where the value of agricultural land is low. Furthermore, searching for a compromise through tree density reduction would reduce costs, without losing significant profits and at the same time, limit the negative impact on agricultural production.

Considering the limitations in comparing the eight scenarios, which involve rotations of 12 and 50 years, to conventional farming, we can expect that agroforestry will ensure a comparative advantage in the future, whereby wood prices are not expected to collapse and the demand for biomass energy will probably grow. The choice of agroforestry practices creates great risks for the farmer, as climatic factors are particularly relevant in determining the capability of forestry species to grow successfully at a specific site. Moreover, in 12 and 50-year scenarios, it is difficult to predict what risks these crops will be challenged by, and additionally, these systems must consider the willingness of the farmers to accept the risk of not seeing an income before 12-50 years, against the typical annual revenue of agricultural crops. In fact, the poor cash flow during most of the rotation could be a demotivating factor for many farmers.

The limitations that emerged from this study are still linked to the lower flexibility that agroforestry guarantees, when compared with annual crops and the lack of awareness of farmers of the opportunities that agroforestry presents. Undoubtedly, agroforestry is able to reduce nitrogen leaching, but in our experimental plots, the timescale was too short to demonstrate this ability. However, from an economic point of view, the results are encouraging, in fact, with a minimum of aid provided by an annual grant, farmers located in NVZ areas could undertake this activity in a profitable way, guaranteeing important environmental benefits.