1 INTRODUCTION

Over the last 20 years, increasing regional and global forest cover has been promoted for a diverse set of reasons, including erosion control, protection of biodiversity, carbon storage and commercial opportunity (Carle, 2006; Dave et al., 2018; Secretariat of the Convention on Biological Diversity, 2014; Zhou, Zhao, & Zhu, 2012). Concern for biodiversity, resilience and ecosystem functions has increasingly governed forestation and reforestation initiatives, promoting the planting of more diverse stands and native species where appropriate (Brockerhoff, Jactel, Parrotta, Quine, & Sayer, 2008; Chazdon, 2008; Galik & Jackson, 2009; Lamb, 2018). In large areas, Mediterranean and Latin America forests are regenerating naturally on marginal and abandoned lands (Bowen, McAlpine, House, & Smith, 2007; García‐Ruiz & Lana‐Renault, 2011) and further commitments to establish forests within landscape restoration initiatives are yet to be enacted, including India's CAMPA fund (Compensatory Afforestation Fund Management And Planning Authority) and a further 80 million ha pledged under the Bonn accord. In this paper, we will refer to all practices as forestation, which we define as a change in land cover from a stable, non‐forested state to a forested one, independent of the long‐term history of forest cover. In order to achieve the ambitious goals that scientists, national and international policies are championing (Chazdon et al., 2016; Griscom et al., 2017; UNFCCC, 2015), a range of forestation strategies will be necessary, appropriate to the specified objective and the national and local context. However, concerns have been raised about the trade‐offs between forest cover and other environmental services, which might be more abundant in a non‐forested state and on which local communities may depend.

There is widespread agreement that forest establishment is associated with a decrease in annual river flow locally, primarily as a result of increasing transpiration and interception rates (Bosch & Hewlett, 1982; Bruijnzeel, 2004; Filoso, Bezerra, Weiss, & Palmer, 2017; Hamilton, 1992; Jackson et al., 2005; Wei et al., 2018). Significant concern has been voiced about the susceptibility of forested catchments to water shortages as a result of changing hydrology (Dymond, Ausseil, Ekanayake, & Kirschbaum, 2012; Jackson et al., 2005). For many years, changing hydrology as a result of forest cover has been studied through its consequences for river flow (Bosch & Hewlett, 1982; Harrold, Brakensiek, McGuinness, Amerman, & Dreibelbis, 1962). River flow is thought to be equal to the difference between precipitation and evapotranspiration and interception, with changes in ground water storage generally discounted at annual timescales (Bari, Smettem, & Sivapalan, 2005; Zhang, Dawes, & Walker, 2001). Whilst water loss from forests to the atmosphere can generate substantial precipitation downwind (Ellison, Futter, & Bishop, 2012; Ellison et al., 2017; Sheil, 2018 although see Angelini et al., 2011), water vapour is unlikely to be entirely recaptured within the same catchment from which it was released, particularly when the catchments are small (van Dijk & Keenan, 2007). Despite the additional role of forests in stimulating rainfall via the release of volatile organic compounds (Pöhlker et al., 2012), forestation is widely reported to result in a net river flow decline for the same catchment (Evaristo & Mcdonnell, 2019; Jackson et al., 2005; Peel, 2009). The impact of forestation on river flow has been reported to increase with mean annual precipitation (MAP) and forested area (Bosch & Hewlett, 1982; Farley, Jobbágy, & Jackson, 2005; Peel, McMahon, & Finlayson, 2010) and more recently with actual evapotranspiration (Evaristo & Mcdonnell, 2019). Larger river flow responses are reported when afforesting grassland rather than shrubland (Farley et al., 2005) and variation in river flow response by forest type (FT) is frequently reported (Farley et al., 2005; Peel et al., 2010; Zhang et al., 2017). Changes in river flow following afforestation are also thought to be more negative than those of reforestation as a result of the added benefits of increase infiltration from establishment on degraded soils (Bruijnzeel, 2004). Areas that were forested historically may have been through substantial transformations in the intervening period, further influencing responses to the reestablishment of forests (Jackson & Hobbs, 2009). These consequences are drawing attention in the context of regional planning and natural capital assessment (Garcia‐Chevesich, Neary, Scott, & Benyon, 2017; Dymond et al., 2012; Jackson et al., 2005; Jones et al., 2017). It is recognized that whether a decrease in river flow constitutes an ecosystem service or disservice is context specific (Dittrich, Ball, Wreford, Moran, & Spray, 2018; Dymond et al., 2012). However, significant questions remain regarding long‐term trends in river flow response to forestation, and how responses will be affected by an increasingly variable climate (Blöschl et al., 2007; Egginton, Beall, & Buttle, 2014; Locatelli & Vignola, 2009; Yan et al., 2019). Although spontaneous forest regeneration may be key to facilitating large increases in forest cover (Yu et al., 2019), considerably less is known about the hydrological consequences of this process (Peel, 2009), despite important work on this question in South Africa and China (Turpie, Marais, & Blignaut, 2008; Yu et al., 2019). It is important to improve our understanding of the impacts of forestation on water supplies over time, under a range of conditions, and as a result, the potential ecosystem service costs to local regions.

The majority of reviews that compare river flow responses to forestation report average changes in flow per catchment (Evaristo & Mcdonnell, 2019; Filoso et al., 2017; Zhang et al., 2017), and few have examined temporal trends at annual or sub‐annual timescales (Farley et al., 2005; Jackson et al., 2005). Whilst the former benefit from a much larger data pool for spatial analyses, information is lost regarding the progression of river flow responses through time. In addition to mean changes in river flow, understanding: (a) how rapidly changes in river flow occur; (b) the magnitude of variation in the system and (c) what the ultimate stable state of the system might be, is key to interpreting how the benefits and costs of forest cover will interact through time (Asbjornsen et al., 2011; Ellison et al., 2017; Farley et al., 2005; Vertessy, Watson, & O'Sullivan, 2001; Vose et al., 2011). This is particularly true in the context of climate change, which is expected to have significant implications for water security across vast regions (IPCC, 2014). Forest water use is known to respond substantially to annual variation in climate (Llorens et al., 2010; Plaut et al., 2013). In arid regions, forests are more likely to be water limited and responsive to pulses in water availability with increased transpiration, whereas humid catchments are more likely to be energy limited (Asbjornsen et al., 2011). Where water is abundant, changes in interception and evaporative demand are responsible for the majority of increased evapotranspiration following forest establishment (Bruijnzeel, 1990). Climate change is expected to result in greater climatic variation around the world and understanding how these changes translate to the catchment scale is important, but it is also important to account for these processes if we are to get a reliable impression of the effects of forest age on catchment hydrology. No prior review of river flow responses through time has separated the consequences of changing climate from those of forest age.

Understanding how catchment hydrology is likely to change in the decades following forestation will be crucial for the sustainability of forestation initiatives. Evidence for reduced forest transpiration with age can be found in maturing pine (Delzon & Loustau, 2005; Mencuccini & Grace, 1996) and eucalypt stands (Haydon, Benyon, & Lewis, 1997; Vertessy et al., 2001) where overstorey transpiration declined by approximately 40%–70% over 45 years from a maximum at approximately 10 years of age. These observations, and age‐related declines in net primary productivity, have been linked to reduced leaf area index, reduced sap wood area index and reduced transpiration per unit leaf area with age (Haydon et al., 1997; Ryan, Binkley, & Fownes, 1997; Vertessy et al., 2001). Interception rates have also been reported to decrease with age as a result of spatial clumping, and reduced leaf area index (Barbier, Balandier, & Gosselin, 2009; Vertessy et al., 2001), although this is highly dependent on species‐specific growth forms (Huber & Iroumé, 2001). These observations are associated with even aged, monospecific stands and it is likely that successional turnover would offset many of these patterns at the catchment scale. However, greater rates of infiltration at intermediate stand densities may also lead to an increase in river flow as naturally regenerating forests develop (Ilstedt et al., 2016). Where forests establish on highly degraded soils, increasing infiltration can significantly influence catchment hydrology (Bruijnzeel, 2004; Wilcox & Huang, 2010). As we use non‐forested land cover as a baseline in this study, we would not expect river flow to ultimately return to baseline levels, but partial recovery may be observed. If partial river flow recovery (hence forth, river flow recovery) following forestation is widespread, it will have substantial implications for the long‐term trade‐offs between carbon storage and water security in newly forested regions. However, uncertainty remains about the generality of prior observations and whether catchment level observations are driven by forest ageing processes. At the catchment scale, partial river flow recovery has been reported two decades after afforestation with pine species, in a global systematic review using a polynomial relationship, with no clear evidence of recovery in eucalypt plantations, attributed to shorter rotation lengths (Farley et al., 2005). Multiple catchments previously analysed in the context of flow recovery were subject to partial deforestation or destruction during the studied time series (Farley et al., 2005; Scott, Prinsloo, Moses, Mehlomakulu, & Simmers, 2000), which is known to lead to long‐lasting increases in river flow (Bosch & Hewlett, 1982) and must be excluded to determine that signals of river flow recovery are due to forest age. Despite the valuable insights provided by Farley et al. (2005) and Jackson et al. (2005) as the first systematic reviews to examine river flow responses to forestation through time, the influence of temporal variation climate was not accounted for, despite widespread directional trends in climate reported over recent decades. We believe that accounting for the role of temporal variation in precipitation and evaporative demand, and separating it from that of forest age, is an important step to understanding the long‐term ramifications of forestation on river flow. As such, further investigation is required to establish whether river flow recovery following forestation is widespread, and the magnitudes of river flow recovery that can be expected as forests age.

In this paper, we systematically review the factors driving river flow responses to forestation through time. For the first time, we explicitly disentangle the effects of forest age from temporal variation in climate, for multiple sites spanning a range of climatic conditions and land use histories. We separate temporal variation in climate at the catchment scale (referred to as temporal or annual variation) from variation in mean catchment climates (referred to as spatial or between catchment variation). We ask: (a) How is the effect of forestation on river flow affected by variable climate and land use history? and (b) What is the long‐term trajectory of river flow in the decades following forest establishment? We hypothesize that increases in annual precipitation will result in larger decreases in annual river flow following forest establishment, as a result of greater rates of transpiration and interception. We expect this effect to vary with catchment aridity, and to be largest in water‐limited catchments where increased transpiration will likely be observed, in addition to increased interception, which will be observed universally. We expect greater annual potential evapotranspiration (PET) to drive larger decreases in river flow after forestation, as a result of greater evaporative demand, and that this effect will be strongest in humid catchments where interception represents a larger proportion of evapotranspiration and forests are less likely to be water limited. We hypothesize that between‐catchment variation in the rate of change in river flow as forests age will be significantly affected by MAP, the percentage of the catchment forested, whether a catchment was historically forested and whether newly forested land was previously used for agriculture. We hypothesize that larger decreases in river flow will be associated with catchments where a larger area is converted to forest, where MAP is larger, where there is no reported history of forest cover and where no agricultural land use was reported. We expect increases in forest cover and MAP to interact, resulting in greater decreases in river flow, as water availability and landscape capacity for evapotranspiration will increase simultaneously. Finally, we expect that the effect of forest age on river flow will be negative, resulting in progressively decreasing river flow until forests reach maturity. We also expect that patterns of partial river flow recovery will be smaller and less frequent than previously reported, after accounting for temporal variation in precipitation and confounding forest management.