Tropical forest soils are an important source and sink of greenhouse gases
(GHGs), with tropical montane forests, in particular, having been poorly studied. The
understanding of this ecosystem function is of vital importance for future
climate change research. In this study, we explored soil fluxes
of carbon dioxide (CO2), methane (CH4), and nitrous oxide
(N2O) in four tropical forest sites located on the western flanks of
the Andes in northern Ecuador. The measurements were carried out during the
dry season from August to September 2018 and along an altitudinal gradient
from 400 to 3010 m a.s.l. (above sea level). During this short-term campaign, our measurements
showed (1) an unusual but marked increase in CO2 emissions at high
altitude, possibly linked to changes in soil pH and/or root biomass, (2) a
consistent atmospheric CH4 sink over all altitudes with high temporal
and spatial variability, and (3) a transition from a net N2O source to
sink along the altitudinal gradient. Our results provide arguments and
insights for future and more detailed studies on tropical montane forests.
Furthermore, they stress the relevance of using altitudinal transects as a
biogeochemical open-air laboratory with a steep in situ environmental gradient
over a limited spatial distance. Although short-term studies of temporal
variations can improve our understanding of the mechanisms behind the
production and consumption of soil GHGs, the inclusion of more rigorous
sampling for forest management events, forest rotation cycles, soil type,
hydrological conditions and drainage status, ground vegetation composition
and cover, soil microclimate, and temporal (seasonality) and spatial
(topographic positions) variability is needed in order to obtain more
reliable estimates of the CO2, CH4, and N2O source/sink
strength of tropical montane forests.
The importance of tropical forests for greenhouse gas budgets
Soils play a vital role in the global greenhouse gas (GHG) budget. Tropical forest soils, in
particular, represent a net sink of carbon (C)
(Pan et al., 2011), but at the same time,
they are the largest natural source of N2O, with an estimated
contribution of 14 %–23 % to the annual global N2O budget
(Werner et al., 2007). In general, soil CO2 is produced
mainly by root respiration, microbial respiration, litter decomposition, and the
oxidation of soil organic matter (Dalal and Allen, 2008).
CH4 is consumed by methanotrophic bacteria
(Jang et al., 2006); however, forest soils
prone to inundation emit CH4 by methanogenic microorganisms (Archaea domain).
N2O is emitted through denitrification or a number of alternative
pathways (e.g., nitrification, nitrifier denitrification,
chemodenitrification, etc.; Butterbach-Bahl et al., 2013; van
Cleemput, 1998; Clough et al., 2017) but can also be consumed during
complete denitrification (Butterbach-Bahl et al., 2013).
Overall, tropical forest soils emit on average 12.1 t CO2-C ha-1 yr-1 (heterotrophic and autotrophic respiration), slightly
less than the net primary productivity (NPP) (12.5 t CO2-C ha-1 yr-1); i.e., the net C sink (belowground and aboveground) of tropical
forests is ∼ 0.4 t CO2-C ha-1 yr-1
(Dalal and
Allen, 2008; Grace et al., 2006). Under aerobic conditions, CH4 fluxes
vary from -0.7 to -30.0 kg CH4-C ha-1 yr-1, with an average
consumption of -3.0 kg CH4-C ha-1 yr-1, while the mean rate of
N2O emissions from tropical forest soils is 3.03 ± 0.52 kg N2O-N ha-1 yr-1 (Dalal and Allen, 2008),
i.e., 2–3 times higher than the mean N2O emissions from temperate forest
soils (1.0 ± 0.36 kg N2O-N ha-1 yr-1;
Chapui-Lardy et
al., 2007; Van Groenigen et al., 2015).
The understanding of the mechanisms and processes underlying GHG flux
variability has greatly improved over the last decades
(Butterbach-Bahl
et al., 2013; Heil et al., 2016; Müller et al., 2015; Sousa Neto et al.,
2011; Su et al., 2019; Teh et al., 2014). However, there is still (1) considerable uncertainty about the overall balances of many ecosystems
(Castaldi
et al., 2013; Heil et al., 2014; Kim et al., 2016b; Pan et al., 2011;
Purbopuspito et al., 2006), (2) a strong imbalance in field observations,
skewed towards the Northern Hemisphere
(Jones et al., 2016;
Montzka et al., 2011), and (3) a bias towards the quantification of emissions
in lowland forests within the tropics
(Müller
et al., 2015; Purbopuspito et al., 2006; Wolf et al., 2011). For instance,
based on a compilation made of CO2, CH4, and N2O fluxes in
South America (Table S1) from 1983 to 2019, there have only been six studies
carried out on tropical montane forests (i.e., > 2000 m a.s.l., above sea level),
while they represent more than 11 % of the world's tropical forests
(Müller
et al., 2015; Teh et al., 2014). In fact, Teh et al. (2014)
and Spahni et al. (2011)
have argued that tropical upland soils are one potentially important source
of CH4 and N2O that has been overlooked in both bottom-up and top-down
emission inventories; their sink/source strength might be comparable to or
greater than their lowland counterparts and, therefore, quantitatively
important in regional and global GHG budgets.
Altitudinal gradients as a biogeochemical open-air laboratory
To further improve our understanding of the role of tropical forest
ecosystems in the global GHG balance, environmental gradients (altitudinal,
latitudinal, etc.) can offer great opportunities to study the influence of
abiotic factors on biogeochemical processes under field conditions
(Bauters
et al., 2017; Jobbágy and Jackson, 2000; Kahmen et al., 2011; Laughlin
and Abella, 2007), which complements the knowledge on short-term responses
from experimental approaches. In the case of altitudinal gradients, these
responses are driven by abiotic variables that covary with elevation,
which, amongst others, creates a distinctly strong climate gradient over a
short spatial distance
(Bubb
et al., 2004; Killeen et al., 2007; Körner, 2007; Myers et al., 2000).
Moreover, since altitudinal gradients reflect long-term adaptations based on
a broad range of factors, they provide valuable insights into the influence
that climate change may have on ecosystem processes
(Malhi et al., 2010). There
is indeed a growing concern regarding the sensitivity of tropical forests to
climate change mainly because species in the tropics have evolved with
narrow thermal tolerances compared to their temperate counterparts; this
makes them particularly vulnerable to changes in global climate
(Fadrique et al., 2018; Perez et
al., 2016). Therefore, the effects of global warming are expected to be
severe in the tropics, and the understanding and integration of the
magnitude of their feedbacks in the Earth system are important to come up
with appropriate forest management options to mitigate climate change
(Bonan, 2008; Li et
al., 2020).
To address these knowledge gaps, we present a pilot study of the
soil–atmosphere exchange of CO2, CH4, and N2O along an
altitudinal gradient in a neotropical montane forest located on the western
flanks of the Andes in northern Ecuador. The sampling campaign took place
from 6 August to 28 September 2018. Four study sites (Fig. S1) were selected: Río Silanche at 400 m a.s.l (hereinafter:
S_400), Milpe at 1100 m a.s.l. (hereinafter: M_1100), El Cedral at 2200 m a.s.l. (hereinafter: C_2200), and
Peribuela at 3010 m a.s.l. (hereinafter: P_3010). Gas samples
were taken using a static flux chamber method once per day per stratum over 2 weeks. Samples of soil were collected once during the whole field
campaign for the analysis of bulk density (ρb), pH, nitrate
(NO3-) and ammonium (NH4+) content, C and nitrogen (N)
concentrations, stable N isotope signatures (δ15N), and soil
texture. Additionally, soil moisture (expressed as water-filled pore space, WFPS) and soil temperature were measured daily. Specifically, we aimed to
determine the magnitude of the soil–atmosphere exchange of CO2,
CH4, and N2O during the dry season. By working along this
altitudinal gradient, we wanted to explore the potential effect of altitude
on the GHG fluxes of the forest soils. Findings from this research could
provide insights for future and more detailed studies on tropical montane
forests.
(a) Soil CO2 (mg C m-2 h-1), (b) CH4 (µg C m-2 h-1), and (c) N2O (µg C m-2 h-1) fluxes per month at Río Silanche (400 m a.s.l.;
S_400), Milpe (1100 m a.s.l.; M_1100), El
Cedral (2200 m a.s.l.; C_2200), and Peribuela (3010 m a.s.l.;
P_3010). Light gray boxplots indicate the fluxes in August
2018, whereas dark gray boxplots indicate the fluxes in September 2018. Light gray
dots in each boxplot represent the measurements taken each day and black
dots the outliers of the respective site. The dotted red line across the boxes
indicates zero net flux.
Average measurements plus or minus the standard deviations (SDs) of
soil CO2, CH4, and N2O fluxes at Río Silanche (400 m a.s.l.; S_400), Milpe (1100 m a.s.l.; M_1100),
El Cedral (2200 m a.s.l.; C_2200), and Peribuela (3010 m a.s.l.; P_3010) per month.
Note: flux values represent the mean of five chambers per site and measurement
week using four-point time series and considering the constraint set to
evaluate linearity in each measurement cycle (R2>0.65).
What did we see in Ecuador?
Across our study sites, P_3010 (the highest stratum)
exhibited the highest soil CO2 emissions (Fig. 1a and Table 1)
probably due to the dominant role of soil pH and shifts in C allocation
patterns. The highest soil pH in water (pHwater) was observed at this site (Table 2),
and under acid conditions, Sitaula et al. (1995) and Persson and Wiren (1989) reported a decrease in CO2 emissions with
decreasing pHwater. On the other hand, although not measured or
estimated in this study, an increase in fine root biomass is expected in
tropical mountain forests compared to lowland forests due to imbalances or
limitations in resource (water and/or nutrients) availability at higher
altitudes (Bauters
et al., 2017; Leuschner et al., 2007). Therefore, the observed increase in
CO2 emissions at P_3010 might be further driven by an
increase in root biomass as the latter has been shown to be positively
correlated with soil respiration
(Han
et al., 2007; Luo and Zhou, 2006a; Reth et al., 2005; Silver et al., 2005).
Physicochemical soil properties of the study areas
Río Silanche (400 m a.s.l.; S_400), Milpe (1100 m a.s.l.; M_1100), El Cedral (2200 m a.s.l.; C_2200), and Peribuela (3010 m a.s.l.; P_3010) at 5 and 20 cm
depth, including mean values plus or minus the standard deviations (SDs) of bulk density
(ρb), porosity, pH in water (pHwater) and KCl suspension
(pHKCl), nitrate (NO3-) and ammonium (NH4+)
concentrations, bulk nitrogen (N) and carbon (C) content, carbon-to-nitrogen
ratio (C/N), and δ15N signatures from samples of soil taken in
August. Similar lowercase letters in superscript and next to some values
within one row and per depth (5 and 20 cm) indicate no significant
difference at P<0.05 between sites (S_400,
M_1100, C_2200, and P_3010).
Note: mean values plus or minus the SD were calculated from soil samples taken
adjacent to each soil chamber (n=5) except for soil texture for which
composites for each site at 5 and 20 cm depth were made from the soil
samples taken from each chamber.
1 Commonly known as Andisol in the United States Department of Agriculture (USDA) soil taxonomy. 2 Expressed per gram
of dry soil. 3 Calculated by dividing C (%) by N (%) for each soil
sample. 4 Expressed relative to the international standard AIR.
In contrast to P_3010, the low CO2 emissions observed
at C_2200 could be attributed to (1) the lower WFPS (Fig. S3),
(2) the lower contents of C and N (Table 2), or (3) the higher bulk density
(Table 2). The lowest soil water content was observed at this site in August
at 5 cm depth, and exactly in this month, the lowest emissions of CO2
were obtained. The low contents of C and N exhibited in C_2200 (indeed, the lowest from all the sites) could also have hampered the
CO2 emissions (Dalal and Allen, 2008;
Luo and Zhou, 2006a; Oertel et al., 2016). Additionally, this site had the
highest soil bulk density (i.e., lowest porosity), which could have led to a
decrease in soil respiration either by a physical impediment for root growth
or by a decrease in soil aeration for microbial activities
(Dilustro et al., 2005;
Luo and Zhou, 2006a, b).
All sites acted as net sinks for CH4 (Fig. 1b and Table 1) (i.e., uptake
of atmospheric CH4 by soils). During the entire field campaign (10 d), only one chamber at one site (S_400) and on a specific
date (08/09/2018) acted as a net source of CH4 (43.2 µg CH4-C m-2 h-1). However, there were no statistical differences between
months, and all sites exhibited indeed a high temporal and spatial
variability.
Only S_400 and M_1100 (both months) (i.e., plots located at the lower locations) acted as net sources of N2O (Fig. 1c and Table 1), whereas the plots located at the highest stratum
(P_3010 & C_2200) showed a general net
N2O consumption during August and September.
The N2O emissions obtained at the lowest strata corroborate the
literature data on lowland tropical forests
(Butterbach-Bahl
et al., 2004, 2013; Koehler et al., 2009) and could be mainly attributed to
the soil water content, temperature, and N availability observed at these
sites (Figs. S2 and S3 and Table 2). Firstly, N2O emissions in tropical
forest soils are predominantly governed by WFPS which influences microbial
activity, soil aeration, and thus the diffusion of N2O out of the soil
(Davidson et al., 2006; Werner et al.,
2007). Secondly, an increase in temperature leads to an increase in soil
respiration and thus to a depletion of oxygen concentrations, which is
indeed a major driver in N2O emissions. In fact, rising temperatures
lead to a positive feedback in microbial metabolism, in which the stimulation
of mineralization and nitrification processes induces an increase in the
availability of substrates for denitrification and thus to an increase in
N2O emissions (Butterbach-Bahl
et al., 2013; Sousa Neto et al., 2011). Finally, the dependency of N2O
emissions on WFPS and temperature is affected by substrate availability
(NO3-). High contents of NO3- give an indication of an
open or “leaky” N cycle with higher rates of mineralization,
nitrification, and thus N2O emissions
(Davidson et al., 2006). Moreover,
NO3- is normally preferred as an electron acceptor over N2O,
and it can also inhibit the rate of N2O consumption to N2
(Dalal and Allen, 2008).
In contrast to the low elevation sites where net N2O emissions were
observed, P_3010 and C_2200 (Fig. 1c and Table 1) presented net consumption (negative values, i.e., fluxes from the
atmosphere to the soil). From 35 valid measurements, only one resulted in net
emission at P_3010 (range: -9.3 to 0.95 µg N2O-N m-2 h-1), whereas from 36 valid measurements, 19 resulted in net
emissions at C_2200 (range: -104.9 to 9.3 µg N2O-N m-2 h-1). Net N2O consumption is often related to
N-limited ecosystems, and it is presumably the cause in our case. At low
NO3- concentrations, atmospheric and/or soil gaseous N2O may
be the only electron acceptor left for denitrification
(Chapui-Lardy et al., 2007;
Goossens et al., 2001). P_3010 had the lowest content of
NO3- along with the lowest soil δ15N (Table 2),
which clearly reflects the shift towards a more closed N cycle at higher
elevations
(Bauters et
al., 2017; Gerschlauer et al., 2019). In fact, studies performed by Teh et
al. (2014)
and Müller et al. (2015)
in the southern Peruvian and Ecuadorian Andes, respectively, related the
decrease in N2O emissions and thus the potential for N2 production
in soils at high elevations to differences in NO3- availability.
Moreover, Wolf et al. (2011) and
Martinson et al. (2013) have
indicated that N availability was (1) a dominant control on N2O fluxes
and (2) inversely proportional to altitude. In addition, the low N2O
fluxes could also be supported by the high content of clay (Table 2) and
CO2 emissions (Fig. 1a) (i.e., development of microsites for N2O
reduction) along with the low soil water content (% of WFPS) (Fig. S3)
(i.e., better diffusion of atmospheric N2O into the soil) and higher
soil pH value (Table 2) (i.e., less severe inhibition of the nitrous oxide
reductase) observed at P_3010
(Chapui-Lardy et al., 2007).
It is important to mention that the region where these measurements were
taken is characterized by a marked seasonality in rainfall. We measured at
the end of the dry season; thus, it is expected that there will be fluctuations in net
fluxes (sources vs sinks) depending on the season. Moreover, although our
limited dataset did not allow us to corroborate the main drivers that
controlled these fluxes, daily measurements like those carried out here
reflect the importance of evaluating short-term variations. As such, the net
N2O consumption with increasing altitude might be overlooked in an
annual analysis, but it is equally important to (1) understand the mechanisms
behind the production and consumption of N2O and (2) have reliable
estimates of the N2O source/sink strength of tropical forests for
regional and even global GHG budgets. Moreover, in order to corroborate the
net consumption observed at high altitudes and improve the understanding of
N2O dynamics in terrestrial ecosystems, disentangling gross N2O
production and consumption at field scale is needed. Although the most
commonly used method to measure N2O fluxes via static chambers only
allows the quantification of net fluxes, stable isotope techniques would
greatly contribute to our mechanistic understanding of gross fluxes. For
instance, enrichment and natural abundance approaches (18O, 15N)
can be used to identify and estimate the contribution of different microbial
processes to N2O production/consumption
(Butterbach-Bahl et al., 2013; Yu
et al., 2020). Nevertheless, (1) the coupling of isotope techniques with
molecular analyses of functional genes is paramount to fully understand the
complexity of the microbial processes present, and (2) the improvement of
measuring techniques for N2O reduction is needed to close N ecosystem
balances (Butterbach-Bahl et al., 2013; Chapui-Lardy
et al., 2007). In fact, microbial composition and diversity, as well as the
presence or absence of important genes (e.g., N2O reductase nosZ I and nosZ II; Van
Groenigen et al., 2015b) can help to detect N2O consumption.
Similarly, analytical techniques such as Raman gas spectroscopy could be
used to detect and quantify N2 fluxes from denitrification
(Frosch et al., 2016), which is indeed a novel and simple
approach compared to previously widely used techniques that may have led to
underestimations (Fang et al., 2015).
Conclusions and future directions
GHG fluxes from tropical montane forests in South America are particularly
scarce with limited spatial coverage and seasonal fluctuation in rainfall
but important for consideration in future field measurements and modeling
research. Overall, we found an unusual but marked increase in CO2
emissions at the highest altitude, which is probably explained by soil pH and root
biomass even though the latter was not measured or estimated. Our CH4
uptake fluxes exhibited a high temporal and spatial variability but
reiterate the role of humid tropical forest soils as CH4 sinks.
Contrary to the net N2O emissions observed in the lowest strata, the
net consumption at higher elevations seems to be quite unique, and it might
reflect the shift towards a more closed N cycle at higher altitudes that was reported
previously in tropical regions. Our results highlight the importance of
short-term variations in N2O fluxes, but it calls for more and broader
studies especially in tropical montane forests, including the impact of
spatial and temporal variability, forest management events and forest
rotation cycles, ground vegetation composition and cover, and soil microclimate
and hydrological conditions, as well as the implementation of isotope
techniques, the coupling of microbial analysis with N2O fluxes, and the
response of tropical forests to current and future changes in N content.
In terms of spatial variation, GHG fluxes may vary between lower slope,
mid-slope, and/or ridge (see Table S1)
(Courtois
et al., 2018; Teh et al., 2014; Wolf et al., 2011, 2012). Fluctuations of
net fluxes can be observed depending on the season and the transition
between them (see Table S1)
(Butterbach-Bahl
et al., 2013; Kim et al., 2016a). Management events (e.g., thinning, clear
cutting, fertilization, draining improvements) and/or forest composition and
growth stage (e.g., young vs mature forest) may influence, e.g., forest
vegetation, soil characteristics, hydrology, and nutrient management among
others and ultimately lead to changes in soil GHG fluxes
(Barrena
et al., 2013; Jauhiainen et al., 2019; Kim et al., 2016a). Moreover, soil
hydrology (runoff, evapotranspiration, soil moisture, etc.) may affect
biogeochemical cycles
(Kim
et al., 2016a). Microbial composition and diversity could be a key to
understand the variability in N2O fluxes (Butterbach-Bahl
et al., 2013). Changes in N content – due to, e.g., urban development and
increasing use of agricultural land – could cause shifts in soil N cycling
and thus CO2, CH4, and N2O fluxes (Koehler et
al., 2012). Besides this, the effects of climate change on tropical regions
(e.g., increases in temperature and CO2 concentrations, as well as
changes in rainfall patterns and drought events) may also affect soil GHG
fluxes. Therefore, a strategic plan must be implemented. Long-term data of
at least 1 or 2 hydrological years is needed, with sampling intervals
covering seasonal fluctuations and being appropriate for the type of land (i.e., spatial variability across different topographical positions). The effect of
N content and climate change in tropical forests could be evaluated using
laboratory (e.g., incubations under controlled conditions) and/or field
experiments (e.g., see Koehler et al., 2009, 2012;
Hall and Matson, 1999; Martinson et al., 2013) with the use of
altitudinal gradients like biogeochemical open-air laboratories. Finally,
although sampling conditions in tropical montane regions can be challenging,
(1) establishing networks and collaborations with local communities (i.e., citizen science) could contribute not only in terms of data acquisition but
also in the development of local knowledge (e.g., how climate and land use
change might affect ecosystems and people), and (2) modeling approaches for C
and N biogeochemistry in forest ecosystems (e.g., Forest-DeNitrification-DeComposition, DNDC, model; GRAMP, 2013) could help to upscale fluxes from site to
regional level. Nevertheless, the cooperation and contribution between field
researchers and scientific organizations, e.g., in South America and around
the world, as well as the capacity building in the respective countries, are
crucial for improving our understanding of soil GHG fluxes from tropical
regions and paramount for getting tangible datasets of remote regions such as
montane forests.
Data availability
The data used in this study have been published in Zenodo and are available under: https://doi.org/10.5281/zenodo.4412827 (last access: 3 Januray 2021, Lamprea Pineda et al., 2021).
The supplement related to this article is available online at: https://doi.org/10.5194/bg-18-413-2021-supplement.
Author contributions
MarB, HV, SeB, and PB developed the project. PALP and MarB
carried out the fieldwork and analyzed the data. MatB and SaB provided technical and analytical support in analyzing the gas and soil
samples. All authors contributed to the ideas presented and edited the
paper.
Competing interests
The authors declare that they have no conflict of interest.
Acknowledgements
We thank Mindo Cloud Forest Foundation, El Cedral Ecolodge, and the Escuela
Politécnica Nacional del Ecuador for the logistic support in Ecuador.
Financial support
This research has been supported by Ghent University and the VLIR-UOS South Initiative COFOREC (EC2018SIN223A103) and COFOREC II (EC2020SIN279A103). Matti Barthel was supported through ETH Zurich core funding provided to Johan Six, and Marijn Bauters is funded as a postdoctoral fellow of the Research Foundation – Flanders
(FWO).
Review statement
This paper was edited by Kees Jan van Groenigen and reviewed by Klaus Butterbach-Bahl and two anonymous referees.
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