**1. Introduction**

### *1.1. Agriculture as Nitrogen Sources*

Nitrogen (N) is an essential element for crop systems, and has been progressively used in agriculture as fertilizers, which made agricultural catchments a major source of N with detrimental effects on the quality of inland and coastal surface waters, including an increased incidence of eutrophication [1–3]. In this context, precipitation and mineralization of the soil organic matter are non-controllable factors which highly regulate the level of N leaching and subsequent loss [1]. Application of fertilizers in agricultural fields, on the other hand, can be regulated at the farm-scale, and anticipates the excess N content in the soil profile prone to leaching [1,4,5]. Therefore, efforts have been made to decrease or optimize the use of fertilizers by balancing N inputs with crop uptake through proper timing and rate of application in order to prevent N accumulation (residual N) and loss [1,6].

Management practices at the farm-scale, often supported by agricultural policies and measures such as the Water Framework Directive, are crucial to reduce N losses and mitigate the impact of agriculture. However, these are often insufficient to reduce N losses to desired levels and protect surface waters [5,7,8]. This normally occurs when the N surplus or storage in the soil profile is particularly high. A common challenge is to find a fine balance between N availability and crop growth through the application of fertilizers so that optimal production and minimal N leaching are achieved [4–6]. Furthermore, the heterogeneous distribution of N across agricultural fields complicates this challenge [6]. Therefore, application of fertilizers above the required level, thus producing N surpluses, are normally the case [4]. In spite of that, crop rotations that include perennial crops have

demonstrated great capacity to reduce N leaching [5,9], although limitations can occur owing to climatic variations [6] or difficulties to make the activity profitable [7]. Moreover, controlled drainage has been used to control the loss of N by regulating the water table level at the site and the resulting outflow [10]. However, this practice can be somewhat complex (e.g., in steep terrains) and hinder crop production [6,10], besides increasing the risk of surface runoff and phosphorus loss due to the emergence of reducing conditions in the soil [11]. Cover crops, in turn, have been reported to reduce N leaching through uptake and storage of organic N in some plant species between crop seasons [6]. Short and cold periods between crop harvest and new planting, however, may limit the effectiveness of this practice. The effect of tillage, on the other hand, can be negligible [1,4] or even negative by supporting N mineralization and subsequent leaching [5–7]. Finally, these practices may also be limited in catchments rich in organic matter, where N mineralization is promoted [1], or in the so-called critical source areas, where N losses are markedly high.

Subsurface drainage has been largely utilized in agricultural catchments with lowpermeable and fine-textured soils (e.g., loamy and clayey soils) to allow proper water infiltration and prevent waterlogging at the site so that agricultural activities can be carried out [12]. However, the drainage networks also function as a direct conduit of nutrients, speeding up the transport of N to surface waters downstream, while minimizing surface runoff and the transport of sediment particles [1,7,10]. Macro-pores in the soil profile and higher soil permeability can also enhance the leaching of N to tile drains and subsequent transport by promoting preferential flow from the soil surface [13,14]. The transport of N occurs according to precipitation events, which control the water flow, and, together with the N content of the soil, regulate the N concentration in tile drains [1,5]. The resulting amount of N lost then refers to the N load in tile drain (Equation (1)). Moreover, subsurface drainage discharge skips natural removal mechanisms for N in the landscape, often resulting in N loads or concentrations sufficiently high to compromise aquatic ecosystems [4,6].

$$\text{N l/load in tile drain } = \text{ water flow} \left(\text{m}^3 \,\text{yr}^{-1}\right) \times \text{N concentration} \left(\text{g } \text{m}^{-3}\right) \tag{1}$$

Tile-drained agriculture can be highly diverse in space and time in relation to agricultural practices, geology, soil type, topography, hydrology and climate, which all contribute to determine the level of N loss between catchments, as well as within the same due to seasonality and annual differences [7,15]. This consequently results in variable N loads and fractions of the N forms transported in tile drains, as these depend on the local catchment characteristics. Nitrogen is transported in the dissolved forms of nitrate (NO3 −) and ammonium (NH4 +), i.e., the bioavailable N forms for crop systems [6], as well as particle-bound or organic N, which is usually transported in low amounts. Ammonium, however, is prone to bind to negatively charged soil particles and become less mobile than NO3 −, or be nitrified, thus converting into NO3 − [7,9]. Therefore, N transport consists mainly of NO3 −, as it is highly mobile and may be generated in situ.

### *1.2. Need to Recover Ecosystem Services at the Edge-of-Field*

Artificial drainage of large water ponded areas has been intensified in the last century in order to allow the expansion and development of agriculture [4,7,12]. This occurred because natural hydrology was often insufficient to lower the water table and promote optimal conditions for crop growth. This process commonly resulted in the conversion of wetlands and peatlands into agricultural fields, which not only disrupted the hydrological regime, but also gradually restricted the natural capacity of the landscape to reduce N loads from highlands to surface waters downstream [4,9,12]. Moreover, mineralization of the pre-existing organic N pool has been favored in response to both water drainage and agricultural practices (e.g., tillage and seasonal vegetation), consequently increasing the potential N content prone to leaching [6]. Thereby, N losses have been increasingly recurrent, especially in soils rich in organic matter.

The reduced capacity of the landscape to decrease N loads, in addition to the complex spatiotemporal dynamics of N losses from tile-drained agriculture and limitations of management practices at the farm-scale described in Section 1.1, have demanded the implementation of strategies at the catchment-scale capable of effectively lowering N loads in tile drains to acceptable levels. Accordingly, construction or restoration of surface-flow systems has been critical in order to re-establish ecosystem services so that significant reduction of N loads from highlands can occur again [4,12]. Therefore, these systems are normally designed to reproduce the N removal mechanisms of natural wetlands. In addition to that, flood control and enhancements in biodiversity are promoted. Surface-flow systems are located at the edge of agricultural catchments and play a fundamental role on increasing the hydraulic residence time (HRT) of the water flow at the outlet of tile drains so that removal of dissolved and any remaining fraction of particle-bound or organic N in water can occur through mechanisms deemed low-cost. The process is primarily promoted by enlarging the area or volume of the surface-flow medium, thus decelerating the water flow, which subsequently favors sedimentation of particles as well as biogeochemical cycling of N between the water, soil, sediments, plants, macro and microorganisms. At this stage, two N removal mechanisms are especially promoted, i.e., assimilation and storage of N into organic forms by the locally existing biota; and denitrification, which depends on carbon availability and anaerobic conditions. Surface-flow systems normally contain hydrophyte plants and hydric soils, which contribute to these mechanisms.

As a result of the above, edge-of-field measures deemed cost-effective are often necessary to intercept the drainage networks and mitigate the effects of N discharge into surface waters [4,6–8]. These measures are, therefore, commonly recommended in critical source areas [9]. It is known that HRT is a key factor regulating N removal [10,12], thus the water flow in tile drains must primarily slow down. Systems that utilize this approach include surface-flow constructed wetlands (SFWs), restored wetlands and drainage ditches, which allow the subsurface drainage discharge to be collected into a basin so that biogeochemical processes ultimately resulting in the removal of N can occur. These systems have been widely used for decades, thus with proven records to suitably reduce the N loads from tile drains.

### *1.3. Aim of the Review*

Among the surface-flow systems mentioned in Section 1.2, SFWs have become the prevalent practice, accounting for a growing body of research in different aspects, as described in Section 2. However, despite the progress, great variability in N treatment within and between SFWs is still observed to date, which often leads to suboptimal performance and uncertain collective effect in watersheds, ultimately complicating estimates of cost-efficiency for planned systems. Therefore, a thorough understanding of this issue is fundamental to ensure that local N removal goals are achieved in the short and long-term. The successful application of SFWs may subsequently minimize interventions in agricultural activities and production. In line with the above, this review aims to (i) describe the SFWs located at the edge of tile-drained agricultural catchments and intended for N removal, (ii) discuss the driving factors behind performance, the causes of variability and related processes, (iii) highlight their strengths and limitations in relation to N treatment, and (iv) suggest plausible outcomes for specific conditions.

This review has no focus on measures that mitigate N losses at the farm-scale, nor on approaches that intensify denitrification at the edge of agricultural catchments (e.g., woodchip bioreactors and riparian buffers), but rather on the use of SFWs as a strategy to increase the HRT of tile drain discharge and promote N removal. Finally, this review avoided citing studies dealing with effluents other than agricultural subsurface or tile drainage, and bases its structure on that found in Mendes [16].

### **2. Variability in the Performance of Surface-Flow Constructed Wetlands**

The past decades have demonstrated that implementation of SFWs at the edge of tile-drained agricultural catchments is a feasible strategy to reduce N loads, especially in critical source areas. This has led to the widespread use of this practice in large national plans—so that a cumulative effect could be achieved—intending to reduce N pollution to coastal and inland surface waters [9,17,18]. Therefore, SFWs are strategically located systems, built in areas where no natural wetland previously existed, and designed to target individual catchments, hence with great potential to cope with large watersheds when used collectively.

The successful implementation of SFWs for this purpose is currently described by numerous studies reporting varying levels of N removal [19–45]. It is generally observed that the performance of these systems largely depends on the N load in the tile drain, climate and SFW design, which all play a role on regulating the system HRT. The varying characteristics between catchments, seasonality and annual differences affect the aforementioned factors, and contribute to the large intra and inter-variability in the performance of SFWs. This consequently complicates estimations for N removal and results in wide variations in area-based N removal rate (mass area−<sup>2</sup> time−1) and efficiency (%) within and between SFWs (Figure 1a,b). Thereby, in order to understand the performance of SFWs, it is fundamental to critically evaluate the influence of hydrochemistry, hydrology and biogeochemistry.
