Deforestation in Southeast Asia is so intense that up to three-quarters of its forests may be lost by the end of the 21st century (Sodhi et al., 2004) and most of the degradation happens because of conversion of forest to croplands and plantations (Wilcove et al., 2013). In Malaysia and Indonesia, more than 16 million hectares of land, mainly from tropical forests but also to a lesser extent other non-profitable agricultural land such as rubber plantations, was cleared for oil palm (OP) (Yan, 2017). Many of the remaining forests are degraded forests, as they have been partially logged to remove specific tree species, and logging activity has caused an increase in forest openings (Houghton, 2012). In 20 % of the world's tropical forests, selective logging occurs, and it is estimated that this accounts for at least half of the anthropogenic greenhouse gas (GHG) emissions from forest degradation (Pearson et al., 2017). Consequently, forest degradation has been recognized as a source of GHG emissions, but little is known of the emissions from the resulting secondary forests, especially from mineral soils in Malaysian Borneo, Sabah. Due to deforestation, fragments of forest remain isolated from each other, which can have consequences for biodiversity and ecosystem function (Ewers et al., 2011).

OP plantations are one of the main causes of deforestation and forest degradation in Southeast Asia (Lee-Cruz et al., 2013; Wilcove et al., 2013), with some disputes about the extent to which industrial plantations are responsible for the loss of old growth and selectively logged forests in Borneo (Gaveau et al., 2016). OP generates the highest yield per hectare of land of any vegetable oil crop. It is used in food products, detergents, soaps, cosmetics, animal feed, and bioenergy and was hence praised as a wonder crop (Sayer et al., 2012). However, OP agriculture is now known to be responsible for soil degradation, loss of soil carbon (C), and reduced soil fertility due to the conversion and management methods (Guillaume et al., 2015; Lee-Cruz et al., 2013). To create an OP plantation, complete deforestation followed by terracing of the land is often the chosen method and not only in hilly terrain. Terracing can result in poor drainage, reduced soil fertility, and increased soil erosion. Conversion of tropical forests also leads to changes in the short- and long-term nutrient status of the converted land-use systems. It is important to understand impacts of these land-use changes in order to identify more environmentally friendly and sustainable management practices (Jackson et al., 2019).

OP plantations are assessed for their GHG emissions, but rarely have emissions from forests and plantations from the same region been reported together, despite the science community calling for flux measurements from forest and converted land simultaneously (van Lent et al., 2015). Much of the focus has been on GHG emissions from tropical forests on peatland and peatland drained for plantations rather than from tropical mineral soils, because of the serious carbon losses when draining the peatlands for crop production. In addition, more attention has been given to carbon fluxes and storage (Germer and Sauerborn, 2008; Hassler et al., 2015) than emissions from the non-CO2 GHG methane (CH4) and nitrous oxide (N2O). Meijide et al. (2020) identified the need to study all three GHGs together in order to assess total emissions from OP plantations. Even though CH4 and N2O are not emitted at the quantity of CO2, their global warming potentials (GWPs) per molecule are 28 and 34 (without and with climate–carbon feedback) and 265 and 298 times higher than CO2 on a 100-year time horizon respectively, which highlights their importance in the climate change debate (Myhre et al., 2013). Due to the serious environmental issues arising from conversion of peatlands to OP plantations, the focus will increasingly shift to mineral soil for conversion to plantations, especially in Malaysia (Shanmugam et al., 2018). However, there are too few measurements reported of N2O emissions from mineral soils in the tropics to draw firm conclusions about the increase in N2O emissions after land-use change from secondary forest to OP (Shanmugam et al., 2018).

Limited measurement and modelling studies have been carried out on N2O emissions from OP plantations (Pardon et al., 2016a, b, 2017), and not in the context of comparing them with other land uses on the same or similar soil type. Similarly, reported CH4 emissions from mineral soils in the tropics (other than from paddy soils) are lacking. Most studies relating land-use change to trace gas emissions have been conducted in South America and not Southeast Asia (Hassler et al., 2015; Veldkamp et al., 2013). An additional caveat of published studies is that most have only been conducted over short periods of time (Hassler et al., 2015). The lack of reliable long-term and multi-year datasets on GHG balances has been recognized (Corre et al., 2014; Courtois et al., 2019). Studies are often associated with high uncertainties (Henders et al., 2015). Nitrogen availability and soil moisture and texture are the main drivers of N2O fluxes in tropical forests and other soil ecosystems (Davidson et al., 2000). As well as agricultural soils, tropical forest soils have been identified as a major source of N2O (Werner et al., 2007), and soil type influences N2O fluxes in the tropics (Dutaur and Verchot, 2007; Sakata et al., 2015). A recent meta-analysis concluded that globally tropical forests emit on average 2 kg N2O-N ha-1 yr-1, and emission rates will significantly increase after land-use change (van Lent et al., 2015). Tropical forest soils are estimated to contribute 28 % to the global CH4 uptake; hence large changes to this sink could alter the accumulation of CH4 in the atmosphere substantially (Dutaur and Verchot, 2007). However, uncertainties are large due to data scarcity. Only one study from Peninsular Malaysia reported that selectively logged forests may be weaker sinks of CH4 and larger sources of N2O than undisturbed tropical rainforest, at least for a short period, because of the increased soil nitrogen availability and soil compaction due to disturbance by heavy machinery (Yashiro et al., 2008).

Forest conversion to OP has shown differences not only in the chemical and physical soil properties, but also in the soil microbial community composition and functional gene diversity (Tripathi et al., 2016). The diversity and abundance of plant communities fundamentally affect the soil microbial community and their function (Eisenhauer, 2016; Tripathi et al., 2016). As of yet, it remains uncertain how conversion from forest to OP impacts microbial communities and their influence on N2O and CH4 fluxes (Kaupper et al., 2019; Tin et al., 2018). Transformation of tropical forest to, for example, OP plantations reduces bacterial abundance initially and alters the community composition but once established may not necessarily result in reduced bacterial richness in the OP soil (Lee-Cruz et al., 2013; Tripathi et al., 2016).

Although the focus of this paper lies on the comparison of soil GHG flux rates (especially for N2O) and their soil chemical and physical properties, we have taken the opportunity to understand the differences in microbial community composition between forests and OP in situ. A previous study has investigated environmental drivers and microbial pathways leading to GHG emissions under controlled laboratory incubations using soils from a subset of the field locations discussed here (Drewer et al., 2020). The aim here was to broadly characterize the microbial communities at the different sites in the different land uses and use the information alongside other measured abiotic factors in mixed models in an attempt to explain the measured fluxes.

The objectives of this study were

to compare GHG emission rates from different land uses,

The following specific hypotheses are included:

N2O fluxes will be larger from OP plantations due to N fertilizer addition compared to tropical forest.

In light of countries committing to reduce and mitigate GHG emissions, e.g. 2015 Paris Agreement (UNFCCC, 2015), it is important to constrain each country's current emission rates, by providing data from measurements rather than relying on model estimates. In this study, we present much needed data of N2O and CH4 fluxes from logged tropical forests and OP plantations on mineral soil as well as their biochemical characteristics and temporal and spatial variability.

This study focussed on comparing GHG fluxes from different land-use types in the tropics. Our data, although not high-frequency measurements, provide a comprehensive insight into the potential impact of converting logged forests to OP plantations on GHG fluxes. The emphasis of this study is on N2O, with auxiliary measurements of CH4 and soil respiration. To date, only four studies published data of N2O emissions from OP plantations on mineral soil in Southeast Asia using the chamber method that included measurements from a time period of longer than 6 months (Skiba et al., 2020). Only one of these studies included measurements in Malaysia (Sakata et al., 2015). Globally tropical forests are the largest natural source of N2O (Werner et al., 2007). Therefore, the question is whether the N input to OP plantations with lower organic matter (TC / TN) content compared to tropical forests will lead to larger N2O emissions than from forests. Although it has been recognized that N2O emissions are induced by N-fertilizer application in OP, when considering annual or long-term emissions from mineral soil, these fertilization patterns may not have a pronounced or clear effect (Kaupper et al., 2019). For example, N-fertilizer-induced N2O fluxes comprised only 6 %–21 % of the annual soil N2O fluxes in OP plantations in Sumatra, Indonesia (Hassler et al., 2017); the rest was due to other natural processes occurring in the soil. Therefore, our study can be considered representative, particularly as measurements were carried out over 2 years. All three land-use types (logged forest, oil palm, and riparian) showed positive N2O fluxes, albeit with a high variability.

On some occasions, our measured fluxes exceeded the range reported by Ishizuka et al. (2005) of N2O emissions from OP plantations on mineral soil in Indonesia, ranging from ∼ 1–29 µg m-2 h-1, by an order of magnitude (maximum measured at 350 µg m-2 h-1). The highest values reported by Ishizuka et al. (2005) were from young plantations, while the lowest fluxes were reported from older plantations. They suggested the low N uptake of young plantations after fertilizer application and the fixation of N by the legume cover crop could be the reason for the high emissions. On the other hand, low emissions from older plantations could result from higher N uptake by the OP and the absence of legume cover. In their study, N2O emissions were mainly determined by soil moisture (Ishizuka et al., 2005), which was not the case here. Mean N2O fluxes from a sandy soil in Malaysia were reported to range from 0.80–3.81 and 1.63–5.34 µg N m-2 h-1 in the wet and dry seasons respectively (Sakata et al., 2015). This was lower than from a sandy loam soil in Indonesia (27.4–89.7 and 6.27–19.1 µg N m-2 h-1 in the wet and dry seasons respectively) (Sakata et al., 2015), indicating the importance of soil texture, provided that management is the same.

Despite the limited number of measurements in OP plantations on mineral soils and the high variability of results, emissions seem to generally be higher in younger OP plantations (Pardon et al., 2016a). This conclusion is not reflected in our data, as OP2 (young) and OP12 (older) plantations showed larger fluxes than the OP7 (medium age) site; although with a lifespan of up to 30 years, all plantations measured in this study can still be regarded as immature. As in our study, Aini et al. (2015) also found no differences in N2O fluxes in the wet and dry months with fluxes ranging from 0.08–53 µg N m-2 h-1. The range of our measured fluxes exceeded those of these previously published studies. However, it is difficult to generalize, as variability appeared to be high in all studies.

Our measured N2O fluxes from the riparian area were similar to those measured in the OP plantation, as soil properties such as bulk density were more similar to OP than logged forest. There is currently a knowledge gap on GHG emissions from riparian areas (Luke et al., 2019), and more studies are needed to evaluate the effectiveness in terms of nutrient retention and potential GHG mitigation of such buffers. A previously published study from Peninsular Malaysia reported mean N2O emission rates from logged tropical forest sites ranging from 17.7–92.0 µg m-2 h-1 N2O-N, which was significantly larger than from their measured unlogged sites (Yashiro et al., 2008). Even though the range of our measured fluxes from logged forest sites was wider, they are broadly of the same order of magnitude (13.9 ± 171 µg m-2 h-1 N2O-N).

As often the case with GHG studies, the variation in the measured GHG fluxes could not be explained with certainty by any of the measured soil parameters. Our sampling frequency was not high enough to investigate, for example, emission rates after fertilizer application in the OP plantations, and besides, this was not the aim of our study. The wide ranges measured for soil mineral N concentrations and N2O fluxes were likely due to the spatial and temporal variability of the fertilizer application, as the slow-release fertilizer bags were randomly placed around the trees, and with time, the fertilizer release rate slowed down. Apart from no strong correlations with single environmental factors, multiple regression and mixed models were only able to explain around 17 % of the variance including multiple measured parameters. However, applying the Bayesian method, the posterior probability density of the effect of land use on N2O flux confirmed that fluxes from the OP plantations were evidently higher than those from the forests (the area of the OP curve does not overlap with the forest curve), with the riparian area being intermediate (mean fluxes 13.9 (95 % CI: -6.3–41.5) µg m-2 h-1 for logged forests, 46.2 (18.4–97.5) µg m-2 h-1 for OP, and 31.8 (-6.3–130.0) µg m-2 h-1 for the riparian area). We therefore confirm our first hypothesis that N2O fluxes are higher from OP than tropical forests.

Agricultural mineral soils such as OP plantation soils can be methane sinks, with uptake rates usually lower than from forest soils (Hassler et al., 2015), which could also be seen in our data with logged forest showing higher uptake rates but at the same time also showing the highest emission rates. However, we did not see the seasonal cycle reported in Hassler et al. (2015) from Indonesia for any of the three land-use types (logged forest, oil palm, and riparian). The lack of seasonal variability seen in our study might be due to the fact that dry and wet seasons are not as pronounced in Sabah as in other tropical regions (Kerdraon et al., 2020) and that temperature is fairly constant throughout the year.

High soil respiration rates (sum of heterotrophic and autotrophic respiration) are considered to be a sign of good soil health, as it reflects the capacity of soil to support soil life including microorganisms and crops. Heterotrophic soil respiration defines the level of microbial activity, soil organic matter content, and its decomposition whilst autotrophic respiration is the metabolism of organic matter by plants. In a recently published study investigating litter decomposition, soil respiration fluxes in Sabah (also in the SAFE area) were higher from forests than OP plantations (Kerdraon et al., 2020). This was also the general trend in our study despite the high variability of all measured fluxes. Litter input in our plots was larger in the logged forest plots and riparian reserve than the OP. Litter decomposition experiments, conducted in Borneo and Panama, revealed that litter input was more important than litter type. This observation stresses the importance of the amount of above-ground litter for soil processes in general, especially in disturbed habitats or forest converted to plantations (Kerdraon et al., 2020).

To further characterize the different land uses and sites within each land use, analyses of soil microbial communities with different assays targeting different microbial components revealed strong influences of soil properties such as pH but also highlighted that fungal and eukaryotic communities were more affected by management and land use than bacteria. Soil pH is known to have an impact on soil microbial community in the tropics (Kaupper et al., 2019; Tripathi et al., 2012), which may explain the very different bacterial communities in logged forest B with the lowest measured pH of all our sites.

Typically, C and N availability and generally soil fertility are known to decrease after deforestation (Allen et al., 2015; Hassler et al., 2015, 2017; Kaupper et al., 2019). This is also reflected in our data (Table 1), where total C and N values in all OP plantations were lower than from forest soils. Nutrient input through litter is higher in the forest than OP plantations and continuously replenished (Guillaume et al., 2015). Therefore, for microorganisms, OP plantations represent a nutrient-deprived environment (Kaupper et al., 2019). Low total C input can also limit the methanotrophic population size and hence limit CH4 uptake (Krause et al., 2012). Lower soil N concentrations in OP soil have also been shown to limit CH4 uptake when compared with forest soil (Hassler et al., 2015). Exactly how shifts in C and N after converting forest to OP may affect microbial processes involved in N2O and CH4 fluxes remains highly uncertain (Kaupper et al., 2019).

Kaupper et al. (2019) have suggested that microbial biodiversity loss occurs soon after clearance and that bacterial diversity may either be resilient to the change or changes cannot be detected after a sufficient recovery period (>8 years) after deforestation. This conclusion is supported by Tin et al. (2018), who reported that the diversity of the bacterial community in a natural forest in the Maliau Basin in Sabah was comparable or even slightly higher in an OP plantation. Conversely, our study implies distinct differences in bacterial, fungal, and eukaryotic community structures between OP plantations and forests. To what extent these differences impact on microbial processes leading to GHG fluxes is hardly known (Kaupper et al., 2019). Despite our data showing effects of land use and soil properties on components of the microbial communities (fungal and eukaryote), including of microbial community, metrics in the GLMM did not help to explain variability in N2O fluxes. Hence, we partially prove our hypothesis that microbial diversity is determined by land use but have to disprove the latter part of the second hypothesis.

It is possible that a more specific focus on relevant functional gene abundances would yield greater predictive ability. Our parallel laboratory investigation, using soils collected from the field study sites reported here, provides a small piece of information on this matter. We concluded that the main contribution to N2O emissions from the logged forests and OP plantations was driven by proteobacterial nirS and AniA-nirK genes from denitrifier and archaeal ammonia oxidizer communities (Drewer et al., 2020). Providing the combined information of soil biochemical reactions with microbial biodiversity may in the future enable better predictions of GHG fluxes. It is vital to understand underlying longer-term processes that ultimately might regulate GHG fluxes to be able to develop GHG mitigation strategies. The conversion of forest to monoculture plantations is a big threat to ecosystem functioning (Tripathi et al., 2016), yet we are still missing data on microbial communities to make accurate predictions of their fate and function.

In an attempt to broadly upscale our findings, we calculated annual soil N2O emission for the Sabah state based on the data from this study (Table 2), together with land cover area estimates (Gaveau et al., 2016). The Sabah scale median N2O emission estimate had increased from 7.6 Mt yr-1 in 1973 to 11.4 Mt yr-1 in 2015. However, this change is small considering the associated uncertainties, demonstrated by the 95 % CI, -3.0–22.3 Mt yr-1 in 1973 and 0.2–28.6 Mt yr-1 in 2015. The changes in land use resulted in small changes of CH4 flux rates over the 42-year period. Median results suggest that Sabah is a sink for CH4 (4 Mt yr-1) throughout the time period presented. There was a slight decrease to the range of our estimate, suggesting that the sink strength will decrease as more land is converted from forest to OP plantations. These estimates, although highly uncertain, highlight the point that the GHG burden of Sabah is likely to increase as a result of land use change from forest to OP plantations and management.

This 2-year field study of bi-monthly measurements demonstrated that N2O emission rates from mineral soils in Sabah were largest from OP plantations, intermediate from a riparian area, and smallest from logged forests. Very large spatial and temporal variability of fluxes and soil chemical and physical properties were encountered at all sites. Mean CH4 fluxes were low with very high variability and showed no clear trend, and the highest range of fluxes was measured in logged forests. Fungal and eukaryotic communities were related to management whilst bacterial community structures were strongly affected by soil pH, which might have masked any management impacts. Mixed models and multiple regression analysis could only explain 17 % of the variation in the measured N2O fluxes, 3 % of the CH4 fluxes, and 25 % of soil respiration, despite the large number of measured abiotic and biotic parameters. This is not uncommon for GHG fluxes but demonstrates that many more studies, ideally at high temporal and spatial resolution, are required to inform on the impact of land use and climate change on GHG fluxes. Scaling up measured N2O and CH4 fluxes to Sabah using land areas for forest and OP implies that since 1973 N2O emissions have increased and CH4 uptake declined, in line with the proportion of OP plantations replacing forest areas. Using the range of measured fluxes with mean and 95 % CI highlights the large uncertainties still associated with these emission estimates, despite having almost 700 individual data points over 2 years. For CH4, the picture is even more uncertain. More studies on N2O and CH4 fluxes from tropical forests and OP plantations on mineral soil are needed to reduce the uncertainty of their emission rates, and especially for experiments deriving N2O emission factors. Furthermore, the impact of current management systems and future potentially more environmentally friendly plantation management needs to be investigated in order to predict how to maintain ecosystem function and biodiversity, which could have a positive impact on reducing GHG emissions.