Paludiculture as a sustainable land use alternative for tropical peatlands: A review (original) (raw)

Introduction

Peatlands are carbon-rich wetlands formed from the reduced decomposition of vegetation biomass due to waterlogged anaerobic conditions (Page and Baird, 2016). Globally, peatlands are distributed across all continents, varying in structure and function driven by local climate, vegetation, hydrology, and geomorphology. In the tropics, peat soils are generally classified as soils possessing approximately 50% carbon by dry weight (Andriesse, 1988; Page and Baird, 2016). Peatlands play an important role in regulating the carbon cycle. Although they cover ~3%, or 4.23 million km2, of Earth's surface land area (Xu et al., 2018b), they store an estimated 700 Pg of soil carbon, twice the amount stored in tropical forest trees (Bonn et al., 2016; Page and Baird, 2016; Yu et al., 2010). Peatlands hold also ~10% of the global freshwater (Joosten and Clarke, 2002), and have important hydrological functions. These include providing 3.83% of all potable water stored in reservoirs (Xu et al., 2018a), attenuating flooding in nearby areas (Lupascu et al., 2020b), contributing to river base flows (Bourgault et al., 2014; Hooijer, 2005), maintaining groundwater levels in superficial aquifers (Hooijer, 2005), and buffering against saltwater intrusion (Hooijer et al., 2012a, Hooijer et al., 2012b; Silvius et al., 2000). Additionally, peatlands also function as producers of biofuel energy, habitats for wildlife, archives of paleo-information, and cultural landscapes possessing aesthetics and spiritual values (Kimmel and Mander, 2010). The protection of peatlands is therefore an urgent priority for climate change mitigation and conservation (Leifeld and Menichetti, 2018; Page et al., 2011).

Peatlands are primarily threatened by industrial monoculture, small-scale agriculture, and forestry expansion that lead to long-term loss of their stored carbon (Lilleskov et al., 2019; Roucoux et al., 2017; Van Asselen et al., 2013; Wijedasa et al., 2018). Approximately 11% (509,000 km2) of global peatlands experienced some degree of degradation due to human activity, the majority of which are concentrated in the tropics (242,000 km2) (Leifeld and Menichetti, 2018). The first stage of establishing peatland agriculture often entails draining peatland to stabilize underlying peat and create dry conditions for plant growth (Comte et al., 2012; Landry and Rochefort, 2012). The lowering of water table exposes peat to oxygen and facilitates previously-inhibited aerobic bacterial decomposition of organic material, releasing vast quantities of stored carbon as CO2 gas (Hooijer et al., 2012a; Wijedasa et al., 2017; Wösten et al., 1997). It is estimated that CO2 emission from drainage in Southeast Asia alone amounts to approximately 1.3–3.1% of emission from global fossil fuel combustion (Hooijer et al., 2010; Wijedasa et al., 2018), and more than 900 g CO2 m−2 year−1 are emitted for every 10 cm of drainage (Couwenberg et al., 2010). Following drainage, high rates of decomposition, consolidation (i.e. settling) and compaction cause the peat surface to subside below its original elevation. Initial subsidence is approximately 1–1.5 m in the first five years post-drainage (Hooijer et al., 2012a; Hoyt et al., 2020), upon which biological decomposition becomes the predominant driver of longer-term subsidence at a rate of 5 cm per year (Andriesse, 1988; Hooijer et al., 2012a; Hoyt et al., 2020; Wösten et al., 1997; Wijedasa et al., 2018). Drainage also increases fire susceptibility and leaching, thereby driving additional carbon loss (Hergoualc'h and Verchot, 2011; Moore et al., 2013; Turetsky et al., 2015; Waddington and Price, 2000). In the tropics, approximately 1.48 Pg CO2 are emitted annually from degraded peatlands through the processes of oxidation, leaching, and biomass burning (Leifeld and Menichetti, 2018), with the amount of carbon lost contingent upon the type of land use change and management practices (Couwenberg et al., 2010; Hergoualc'h and Verchot, 2011; Murdiyarso et al., 2010).

Presently the conversion of peat swamp forest (PSF) for economic development is especially accentuated in Southeast Asia (Gaveau et al., 2019; Miettinen et al., 2012; Wijedasa et al., 2018). Satellite estimates indicate that 50% of forested peatlands in Peninsular Malaysia, Sumatra, and Borneo had been converted to industrial plantations and other forms of managed systems such as smallholder agriculture by 2015; a stark contrast against 1990 where only 11% of peatlands were converted (Miettinen et al., 2016; Wijedasa et al., 2018). Between 1990 and 2010, conversion of PSF to smallholder agriculture and oil palm plantations in Southeast Asia released 1.46–6.43 Pg of CO2, with another projected 4.43–11.45 Pg CO2 emitted by 2030 under current development trajectories (Wijedasa et al., 2018). In other parts of the tropics such as South America and Africa, newly-discovered peatlands are being threatened by mining and hydropower dam operations, oil and gas extraction, road construction, habitat fragmentation, and climate change (Dargie et al., 2019; Draper et al., 2014; Lilleskov et al., 2019; Murdiyarso et al., 2019).

The role of peatland carbon storage in climate change mitigation has led to an interest in managing peatlands sustainably to protect underlying carbon stock (Hermanns et al., 2017). This include efforts to restore and rehabilitate degraded or abandoned peatlands through revegetation (Dohong et al., 2018; Hytönen et al., 2018). An oft-proposed method is to manage the land under paludiculture (Fig. 1). Paludiculture is loosely defined as the “sustainable production of biomass on wet and rewetted peatlands” (Wichtmann and Joosten, 2007), where biomass refers to any form of material derived from biological origins. Although the use of wet peatland resources dates back many decades, the application of the term ‘paludiculture’ is a recent phenomenon. Given the importance of peatlands as carbon sinks, sustainably using peatlands entails maintaining a neutral carbon balance, with long-term prospects of turning these ecosystems into carbon-negative sinks (Wichtmann and Wichmann, 2011). This goal is reflected in the adoption of paludiculture in climate change-related policies by organizations such as the Ramsar Convention on Wetlands and FAO (Wichtmann and Couwenberg, 2013).

Despite its expanding usage, paludiculture as a concept lacks clear guiding principles. The literature on sustainable agriculture has long recognized the ambiguities in defining broad concepts, such as sustainability in agriculture, that vary based on the environmental, socio-economic, and political context (Pannell and Schilizzi, 1999). Such variations can promote debates and highlight synergies between different schools of thought, which is key to advancing a concept (Velten et al., 2015). Nevertheless, concerns have been raised about the misappropriation of paludiculture to peatland-degrading practices such as partial drainage or higher water table systems with net positive carbon balance, thus eroding the ecological integrity and long-term existence of peatlands (Budiman et al., 2020; Giesen, 2013; Giesen and Sari, 2018; Sari et al., 2018; Tata, 2019; Taylor et al., 2019). It is through critical examination of the common definitions and application of paludiculture that we can unravel perceptions and priorities about the concept and assess how these definitions influence practice and research.

This review aims to answer the following questions: what is paludiculture and what are our current knowledge regarding the potential application of paludiculture in the tropics? We examine existing paludiculture definitions to derive key themes and highlight contentious points. We then review the current literature about peatlands and paludiculture within the framework of the key themes. Due to a scarcity of paludiculture research from the tropics, we draw upon research conducted in northern peatlands – which have a longer history of paludiculture development and potential influence on current definitions – where relevant, to identify areas of research gaps for tropical paludiculture. Finally, we provide recommendations for future paludiculture research in the tropics.

The development of paludiculture is grounded in research from northern peatlands. Biophysical, meteorological, and socioeconomic differences between ecoregions makes it challenging to transfer a concept from northern latitudes to the tropics. Understanding the commonalities and disparities between northern and tropical peatlands is a crucial first step for determining the extent that paludiculture is applicable to the tropics and future research priorities (Table 1).

Peatlands, in general, share several characteristics: they are wetland ecosystems with more carbon input than output and act as long-term carbon sinks, the function of which can be rapidly reversed by disruptions to the hydrology and surface vegetation (Joosten and Clarke, 2002). Although Bacon et al. (2017) argued that peatlands are governed by the same processes at large between the north and the tropics, there exist clear distinctions between the ecoregions. The majority of the world's peatlands occur in the northern hemisphere, encompassing both boreal and temperate regions. In low relief (i.e. poor draining) environments, such as those found in Europe, North America, and Russia, peatlands form under conditions of high precipitation and low temperature (Page et al., 2009). Both boreal and temperate peatlands are predominantly formed from bryophytes and graminoids, displaying slower rates of accumulation and decomposition relative to the tropics (Frolking et al., 2001; Wieder and Vitt, 2006). Global estimates indicate that northern peatlands cover 3,794,000 km2, but only store 448.9 Gt C at a density of approximately 118,318 t C km−2 (Leifeld and Menichetti, 2018). Despite majority of peatland area being found in the boreal ecoregion, paludiculture research is concentrated on temperate peatlands. Hence, for the rest of this review, ‘northern peatlands’ will be used synonymously with temperate peatlands.

Tropical peatlands are mostly found in the lowlands, often located between river bodies and near the coast. While some higher altitude tropical peatlands do exist, their conversion to agriculture is limited. Tropical peatlands are dominated by woody trees with greater primary productivity, resulting in rapid rates of peat soil formation (Wüst et al., 2007). This process is contrasted by accelerated decomposition rates under a high temperature climate, causing degraded tropical peatlands to contribute more substantially to global GHGs emissions (Page and Baird, 2016; Page et al., 2009). Although tropical peatlands cover only 587,000 km2, they store 119.2 Gt C at a density per unit area of 203,066 t C km−2 (Leifeld and Menichetti, 2018).

The different origin of peat-forming plants between northern (i.e. moss/sedge) and tropical peatlands (i.e. woody trees) further results in variations in peat characteristics and responses to disturbances. Tropical PSF trees possess extensive root structures creating large pore spaces, leading to higher hydraulic conductivity and lower bulk density near the surface, and greater water retention capacity in depths of more than 50 cm (Page et al., 2009; Wösten and Ritzema, 2001). Similar patterns of hydraulic conductivity and bulk density are also found in northern peatlands, albeit in smaller magnitudes due to smaller pore spaces. On average tropical peat bulk density can exceed 0.2 g cm−3 for sapric peat (FAO, n.d.) and hydraulic conductivity varies between 0.001 and 13.9 m day−1 depending on land use (Kurnianto et al., 2019), whereas the bulk density and unsaturated hydraulic conductivity of northern peatlands are 0.02–0.254 g cm−3 and 0.07–1.04 m day−1, respectively (Rezanezhad et al., 2016). When disturbed, positive feedback between drainage, tree-clearing, and fires induce vegetation composition to shift towards more fire-prone grasses and flood-tolerant ferns in the tropics, thereby exacerbating disturbance risks (Hoscilo et al., 2011; Miettinen et al., 2013; Page et al., 2009). Vegetation shift towards vascular plants as opposed to bryophytes due to drainage and peat mining disturbances is observed in northern peatlands (Girard et al., 2002; Lachance and Lavoie, 2004). These dynamics pose challenges for restoring peatlands and their ecological functions: Estimates indicate that peat accumulation may be restored after 20 years in bryophyte-dominated temperate peatlands (Graf et al., 2012; Lucchese et al., 2010), whereas the process has not been studied and could take longer in the tropics depending on tree species (Budiman et al., 2020; Harrison et al., 2020).

Another unique characteristic of tropical peatlands is the complexity of PSFs and associated biodiversity. Over 1400 plant species (Giesen, 2013; Giesen et al., 2018; Posa et al., 2011) and 700 fauna species (Posa et al., 2011) have been documented in Southeast Asia PSF alone, compared to more depauperate, but specialized species found in northern peatlands (Joosten and Clarke, 2002; Warner and Asada, 2006). For example, a review of Canadian peatland plant species yielded 711 species, with the majority being herbaceous species and bryophytes (Warner and Asada, 2006). Surveys of a temperate Himalayan peatland found 460 plant species and 165 bird species, some of which are listed as threatened (O'Neill et al., 2020). In comparison, much of the ecology and diversity of tropical peatlands remains understudied, particularly for recently discovered areas in the Congo Basin (Dargie et al., 2019; Xu et al., 2018b) and South America (Gumbricht et al., 2017; Lähteenoja and Page, 2011; Lilleskov et al., 2019; Murdiyarso et al., 2019; Roucoux et al., 2017).

Aside from biophysical differences, the development of peatlands in different ecoregions has followed slightly disparate pathways. Peat mining for biofuel production and livestock grazing are documented extensively in northern ecoregions. In places such as Eastern Europe, large areas of peatlands have been abandoned after the peat extraction industry declined (Joosten et al., 2012). Currently, agricultural use of peatland is ongoing in Europe, North America, and Southeast Asia. Unlike its northern counterparts, tropical peatlands have attracted concerns for the extent, scale and speed by which they are converted to industrial monoculture, especially in Southeast Asia (Dohong et al., 2017; Joosten et al., 2012; Miettinen et al., 2012). Large-scale biomass production such as oil palm (Elaeis guineensis Jacq.) and pulpwood plantations has received the bulk of attention due to their destructive impacts on the carbon stock of peatlands (Evans et al., 2019; Hooijer et al., 2010; Miettinen et al., 2012; Murdiyarso et al., 2019). Moreover, unlike the low population density of northern areas, relatively high population density in parts of the tropics may result in land use competition against paludiculture (Lilleskov et al., 2019). Practitioners should take these unique characteristics of tropical PSF into consideration to guide paludiculture projects and research.