Boreal peatlands store an enormous pool of soil carbon that is dependent upon – and vulnerable to changes in – climate, as well as plant community composition. However, how nutrient availability affects the effects of climate and vegetation change on ecosystem processes in these nutrient-poor ecosystems remains unclear. Here we show that although warming promoted higher CH4 emissions, the concurrent addition of N counteracted most (79%) of this effect. The regulation effects of the vegetation functional group, associated with the substrate quality, suggest that CH4 emissions from peatlands under future warming will be less than expected with predicted shrub expansion. In contrast, N2O flux will be enhanced under future warming with predicted shrub expansion. Our study suggests that changes in greenhouse gas emissions in response to future warming and shifts in plant community composition depend on N availability, which reveals the complex interactions that occur when N is not a limiting nutrient.
Northern peatlands store ~30% (~600 Gt) of the world’s terrestrial soil carbon (C)1, equivalent to half of the total atmospheric C2. This enormous store of soil C results from persistently greater rates of plant production than decomposition, due to the high water content, poor nutrient3, and recalcitrant litter such as Sphagnum moss4, all of which reduce decomposition. However, the anoxic conditions of northern peatlands make them a global source of methane (CH4), annually releasing 10–25 CH4-C Tg (12.2% of the global total) into the atmosphere5.
Carbon dioxide (CO2), CH4, and nitrous oxide (N2O) are the three most important greenhouse gases (GHGs), after water vapor. Emission of GHGs by northern peatlands are tightly coupled to climate change through the impact of climate on peatland hydrology and plant community composition6,7. One concerning factor is that the majority of peatlands are located in northern high latitudes where the climate is experiencing a greater rate of change than in the past2, and climate warming is expected to increase nutrient mineralization from soil organic matter8. Moreover, the C pool in northern peatlands is susceptible to changes in climate6 via changes in temperature, soil water content, and soil nutrients. Further, climate warming can alter the vegetation composition, e.g., shrub expansion in tundra areas9,10,11, shifting the dominance of plants from Sphagnum to a graminoid-dominated system in poor fens12, or leading to the loss of selective plant species9,11,13,14. Land use change also affects the plant community composition. For example, burning and grazing promoted fast-growing graminoids over slower-growing ericaceous shrubs and mosses15, while drainage reduced the coverage of Sphagnum moss on hummocks that facilitated the invasion of sedges on lawns in a poor fen16. Previous studies have shown the rapid response of C17,18, even subsurface peat19 or methane emission17, to manipulative warming from peatlands. The change in vegetation composition or biodiversity loss can also exert severe impact on both short-term C fluxes20 and long-term soil C storage21,22,23. For example, the presence of graminoids (sedge-dominated in this case) has previously been demonstrated to positively impact CH4 flux, either by facilitated transportation due to the presence of aerenchymatous tissues16,24,25 or by increased supply of available substrates for methanogenic activities26,27. The complex interactive effects between abiotic and biotic variables on ecosystem C processes are becoming more apparent28. It was illustrated that the effect of warming on GHG fluxes in peatlands are modulated by plant community composition25.
Northern peatlands tend to be nutrient-limited with slow rates of decomposition3. Carbon storage in boreal ecosystems is thought to be constrained ultimately by C–nutrient interactions because plant production is usually nitrogen (N)-limited29. It has been found that northern peatlands with different soil N concentrations show very different responses to increases in temperature30. Different vegetation types, for example shrubs, sedges, and Sphagnum mosses, have been illustrated to show disparate responses to experimental N addition31,32, e.g., increased above-ground vascular plant biomass (e.g., Vaccinium oxycoccus33), reduced peat-forming Sphagnum32, or change in species composition34. Therefore, we hypothesized that the combined effects of warming and vegetation shifting on ecosystem processes largely depend on N availability. Unraveling the underlying mechanism is crucial because the global N deposition is predicted to double by 205035.
Here, we report the results from a manipulative experiment based on a fully factorial design enabling us to examine the interactive effects of passive warming and plant community composition on GHG emissions under both N-ambient and N-added conditions from a boreal peatland (Supplementary Figure 1 for the experimental design). We expected the system would experience a maximum response in the short term as we were suddenly inducing a disequilibrium. Further, we acknowledged that the short-term responses would be transient until a new equilibrium is reached in the long run. However, we argue that the transient results would be still very useful because they would offer us the information of whether the system would arrive at the same equilibrium or move to a new equilibrium in the long run. Hexagonal open-top chambers36 were placed on half of the experiment plots in an area of oligotrophic blanket bog in Newfoundland, Canada, to achieve an ~1.2 ℃ increase in soil temperature at 5 cm depth during the mid-day period. Warmed or ambient temperature plots were manipulated by both the addition of N and the removal of selected vegetation types, including graminoids and shrubs, while the moss layer was kept intact to minimise the soil disturbance. The experiment was carried out for 2 years. We present results from field measurements of GHG fluxes, namely CO2 (here represented by ecosystem respiration), CH4, and N2O fluxes during the second growing season. Our results demonstrate that the increase in CH4 emissions from northern peatlands in response to climate warming may be substantially smaller than previously predicted with elevated N deposition; the projected shift to increased shrub cover in boreal peatlands may lead to a less pronounced response of CH4 emissions to climate warming, but a stronger N2O exchange between the atmosphere and peatland ecosystems under the projected warming along with increasing atmospheric N inputs.
Open-top chambers enhance air temperature
Open-top chambers statistically significantly increased soil temperature, on average, by 1.2 °C (F1,384 = 18.2, P < 0.001) at 5 cm depth and 0.44 °C (F1,384 = 29.8, P < 0.001) at 20 cm depth (Supplementary Tables 1 and 2; Supplementary Figure 1). We found no evidence that either N (with addition of 6.4 g N m−2 year−1; F1,384 = 0.05, P = 0.83) or vegetation manipulation (graminoid present/absent: F1,384 = 1.06, P = 0.31; shrub present/absent: F1,384 = 0.07, P = 0.79) affected soil temperature. All treatments of our experiment had no significant effects on water table depth (Supplementary Tables 1 and 2).
Nitrogen addition, warming, and vegetation removal had no significant effects on CO2 flux (Fig. 1a, Table 1). Although warming increased CO2 flux by 25%, from 162.1 ± 13.4 (SE) to 203.3 ± 14.1 mg m−2 h−1 for sites with both shrub and graminoid present (G + S), this increase is statistically not significant (F1,83 = 3.39, P = 0.06). A consistent seasonal pattern of CO2 flux among treatments was observed (Supplementary Figure 2), but there were no interactions between sampling time and treatments (Supplementary Table 3).
N addition counteracts warming-induced CH4 emission
Our analysis revealed significant treatment effects from both warming and plant community composition on CH4 emission (Fig. 1b, Table 1). Specifically, the warming doubled the overall mean CH4 emission from 1.39 ± 0.29 to 2.85 ± 0.39 mg m−2 h−1 regardless vegetation composition, and CH4 emissions were larger in the presence of graminoids (2.24 ± 0.23 mg m−2 h−1) than without graminoids (1.46 ± 0.21 mg m−2 h−1) independent of other treatments. Although N addition alone did not influence CH4 emission (F1,374 = 1.51, P = 0.23), we observed a significant combined effect of N addition and warming on CH4 emission (F1,374 = 9.27, P = 0.004). More specifically, N addition counteracted most (~79%) of the warming-induced increase in CH4 emissions (Fig. 1b) and led to an overall mean CH4 emissions rate of only 1.69 ± 0.27 mg m−2 h−1 (a 0.30 mg m−2 h−1 increase) for combined treatments (compared with 2.85 ± 0.39 mg m−2 h−1 for the plots receiving the treatment of warming independent of vegetation composition). On the other hand, we detected an interactive effect on CH4 emissions of warming with the removal of graminoid (F1,374 = 10.1, P = 0.003), with the lowest CH4 emission rates being measured in unwarmed plots without the graminoids (0.16 ± 0.47 mg m−2 h−1, a greater than 0.8 factor decrease) and the highest CH4 emission rates in warmed plots with intact vegetation (3.73 ± 0.87 mg m−2 h−1, Fig. 1b). Additionally, the presence of graminoids or shrubs interacted to affect CH4 flux (F1,374 = 4.71, P = 0.04). No interactions between N addition and vegetation manipulation were found to affect CH4 emissions (Table 1).
Warming and shrub expansion enhances N2O flux
The N addition increased the N2O flux by 44-fold, from an overall mean of 0.011 ± 0.026 mg m−2 h−1 for the control treatment to 0.48 ± 0.11 mg m−2 h−1 for the N addition treatment (F1,304 = 22.8, P = 0.001; Table 1 and Fig. 1c). We did not detect a significant effect of warming alone on N2O flux (F1,304 = 0.17, P = 0.69). An interactive effect was observed between N addition and warming (F1,304 = 6.36, P = 0.03), wherein warming further enhanced the positive effect of N addition on N2O flux by 67% (Fig. 1c). The interactive effect between N addition and warming varied according to plant functional groups (Table 1), with the largest N2O flux being found at sites with the treatments of N addition and warming where only shrubs were present (1.85 mg N2O m−2 h−1) and the smallest at sites with the treatments of N addition and warming where only graminoids were present (0.17 mg N2O m−2 h−1, Fig. 1c). Significant interactions between warming and graminoids (F1,304 = 11.4, P = 0.007) and between warming and shrubs (F1,304 = 15.6, P = 0.003) were detected, suggesting the effect of warming on N2O flux was regulated by the removal of different plant functional groups (Table 1, Fig. 1c). The presence of graminoids or shrubs, without N addition and warming, also interacted to affect N2O flux (F1,304 = 5.93, P = 0.035).
Vegetation regulates substrate quality altered by warming
Warming significantly increased pore water dissolved organic C (DOC) from 35.7 ± 1.40 to 37.3 ± 1.01 mg L−1 (P = 0.04) (Table 2 and Fig. 2). N addition did not significantly change DOC (Table 2). A significantly higher DOC was observed in plots where shrubs were absent than in those where they were present (F1,382 = 3.48, P = 0.07, Table 2). Significant interactions were observed among warming, N addition, and graminoid removal on DOC (Table 2). Warming slightly increased DOC aromaticity, where peatland derived DOC with high aromaticity is often linked to low bioavailability37, indicated by specific UV absorbance at 254 nm, SUVA254 (P = 0.07), and graminoid removal significantly increased DOC weight-averaged molecular weight (indicated by a UV absorbance ratio between 250 and 365 nm, a250/a365, F1,317 = 6.69, P = 0.01, Table 2, Fig. 3). Our data indicated that N addition and warming interactively affected DOC, TN, C/N ratio, and the chemical composition of DOC with the effects being regulated by vegetation composition.
In our study, we manipulated both abiotic (temperature, N level) and biotic (presence or absence of vascular plant functional groups, i.e., shrubs and graminoids) factors to investigate the complex interactions regulating ecosystem processes, i.e., ecosystem respiration (Reco), CH4, and N2O fluxes. Our findings agree with our hypothesis that soil N condition regulates the independent, or the combined effects of warming and vegetation shifting on ecosystem processes. Based on our findings, we generated a conceptual model of GHG emission by northern peatlands (Fig. 4), which presents the complex regulating effects of soil temperature, soil moisture or water table depth, N conditions, and vegetation composition on GHG emissions. Particularly, this model emphasizes the interactive effects between warming and N deposition on CH4 and N2O fluxes, and between vegetation composition and N deposition on N2O flux, which were revealed for the first time through this comprehensively manipulative experiment that considered changes in climate, N availability, and vegetation at the same time.
The insignificant change of Reco could be attributed to the slow stabilization of plant community composition to vegetation removal because we observed that warming slightly increased Reco at the plots without vegetation removal, but this increase was not statistically significant (P = 0.06). Since our experiments have only run for 2 years, our findings must be viewed with caution. We still need to see if this insignificant change in Reco would persist after a long-term period of warming and vegetation shifting (i.e., after 5 years’ treatment).
In contrast, the two potent greenhouse gases, CH4 and N2O, showed strong responses to the manipulations. Considering first CH4, warming unsurprisingly overall increased CH4 emissions by 105% regardless of vegetation treatment, which was consistent with previous research on peatlands25,38. The increased CH4 emissions resulted from warming may be due to increased methanogen activity, and substrate quantity (Table 2 and Figs. 2, 3) as a consequence of stimulated vegetation growth39. Previous studies have been contradictory regarding the impact of N addition on CH4 emissions, which have been shown as either a significant increase40 or a statistically insignificant change41,42. This has led to conflicting reports on the net impacts of N fertilizers on CH4 emissions43,44. Therefore, the insignificant effect of N addition on CH4 flux (Table 1) was not unexpected. However, N addition strikingly counteracted most (79%) of the warming-induced CH4 emissions (Fig. 1b), implying that future warming may not trigger as large a release of CH4 emissions from peatlands as has been anticipated38,45 in the presence of increasing N deposition from the atmosphere. Note must be made that the counteracting effect in our experiments could largely depend on the non-limited N supply in our plots as we used 10 times higher N than reported N deposition values in the N addition plots.
The mechanisms underlying the above counteracting effects may be two-fold. Given that the balance of methane production and consumption determines the net flux of CH446, the suppression of methane production by nitrate47,48 could be enhanced by warming. Further, the methane oxidation by nitrogenous fertilizers43,49 may be stimulated under warmer conditions, which is less likely because CH4 oxidation rates in soils are theoretically less temperature dependent than CH4 production50. We also speculate that an increased supply of nitrite (NO2−) produced by nitrification or denitrification processes combined with a warmer climate would accelerate the process of anaerobic methane oxidation driven by oxygenic bacteria51. No matter which mechanism (if any) underlies the counteractive effect (Fig. 4), the negligence or unawareness on this issue52 leads to an over-estimation of future CH4 emission in response to warming based on our current understanding53. Caution has to be made that this is under the consumption of an unchanged water table since the influence of temperature on CH4 emissions is hydrologically dependent54. The regulating mechanism urgently needs to be disentangled through investigation of soil microbial dynamics under the conditions of warming and N addition.
The presence of graminoids (sedge-dominated in this case) has previously been demonstrated to positively impact CH4 emission, either by facilitated transportation due to the presence of aerenchymatous tissues16,24,25 or by increased supply of available substrates26,27. Consistently, graminoid removal that increased high molecular weight DOC and aromaticity (Table 2 and Figs. 2, 3) in our study may inhibit CH4 production. Here we observed that warming enhanced the facilitation effect of graminoids, suggesting that the direction of shifting in plant community composition induced by climate change14 determines CH4 emission. For example, the predicted shrub expansion in peatlands due to warming11 may interactively retard the warming-induced CH4 emissions. The regulation of substrate quality by a plant functional group may contribute to the interactive effect of warming and vegetation removal on CH4 emissions, since the sensitivity of organic matter decomposition to warming varies with substrate quality55.
N2O flux in our study was relatively low (0.011 ± 0.026 mg m−2 h−1), as would be expected in this nutrient-poor ecosystem, within the range (0.0004–0.034 mg N2O m−2 h−1) of previous reports from boreal peatlands56. Interestingly, warming alone did not affect the N2O flux, while interactively enhanced the effects of either N addition or functional group removal on N2O flux. In contrast with CH4, boreal peatlands would act as a stronger N2O source under future warming conditions with predicted shrub expansion10,11, while the N2O sink of sedge-dominated peatlands may be enhanced under future warming. These interactions (Fig. 4) have rarely been examined by field experiments or process-based modeling, which could lead to an over or underestimate of N cycling for northern peatlands that are being or will be dominated by specific plant functional groups under the condition of elevated N deposition. Although seasonal variations were observed for GHG fluxes (Supplementary Figure 2), no interactive effects of sampling time with treatments (Supplementary Table 3) were found in our data, which suggests that our treatment effects on GHG fluxes are independent of the measurement time.
In summary, our findings provide direct evidence that interactions occur among warming, N addition, and plant community composition to modify the GHG emissions from boreal peatlands. To the best of our knowledge, this is the first time that a counteractive effect of N addition on warming-induced CH4 emissions has been observed. If such a damping effect of N-related suppression of methane production was to occur globally, the increase in CH4 emissions from northern peatlands in response to climate warming may be significantly smaller than previously predicted, and thus may not cause as much positive climate feedback as anticipated. However, the universality and the experimental results need to be verified in different locations before the findings can be extrapolated to other peatlands. Moreover, the projected shift to increased shrub cover leads to a less pronounced response of CH4 emissions to climate change than expected by the decrease in substrate quality. In contrast, the projected shift to shrub cover may lead to stronger positive responses in terms of N2O exchange between the atmosphere and peatland ecosystems under predicted warming along with increasing atmospheric N inputs. We emphasize an urgent need to unravel the underlying mechanisms before incorporating the interactions between biotic and abiotic drivers into future modeling work. Our findings, for the first time, show a decisive control of soil N condition on the independent and interactive effects of warming and plant community composition on ecosystem processes. Moreover, our experiment can serve as a possible direction to ecosystem management in terms of mitigating climate change.
Our research site is located in an area of oligogenic, ombrotrophic blanket bog, in Robinsons, western Newfoundland, Canada (48°15′46′N, 58°39′ 21′W). The climate is oceanic temperate, with an annual rainfall of 1340 mm and annual average temperature of 5 °C (1981–2010)57. The mean pH (1:5 soil/water) at the site was 4.5 ± 0.01, and the mean peat depth of 3 m was derived from three random peat depth measurements at the site before the experiment was established. The site represents the typical type of peatland found on the island of Newfoundland, where the vegetation consists of an approximately equal biomass of graminoids (Trichophorum cespitosum, Carex chordorrhiza) and dwarf shrubs (Gaylussacia baccata, Rhododendron groenlandicum, Andromeda glaucophylla, Ledum palustre ssp.), with Sphagnum mosses (Sphagnum spp., Hylocomium splendens, Aulacomnium turgidum) providing the main matrix39,57.
A factorial design comprising manipulation of temperature (warmed vs. ambient temperature), nitrogen (N; N addition vs. no N addition), and plant community composition (removal of graminoids only vs. removal of shrubs only vs. removal of both graminoids and shrubs vs. no vegetation removed) was established in the spring of 2014. The layout was 2 (Warming and ambient temperature) × 2 (N addition and no N addition) × 2 (shrubs present or absent) × 2 (graminoids present or absent) = 16 treatments. The experimental site comprised four blocks with 6 m apart from each other, and each block contained one each of all 16 treatments, randomly distributed (n = 64) (See Supplementary Figure 1 for the layout of the experiment design) 2 m × 2 m plot, with a buffer zone of intact vegetation of at least 2 m between adjoining plots.
Warming was achieved passively using hexagonal open-top chambers based on the ITEX design58. Each transparent section making up the hexagonal open-top chambers measured 80 cm along the bottom edge, 62.5 cm along the top edge, and 40 cm in height, to give an internal diameter of 1 m2. The chambers were constructed from 3-mm thick clear acrylic sheeting (Ridout Plastics, San Diego, U.S.), which allows 92% light transmission. The open-top chambers method offers a robust means to examine the effects of warming in remote environments without the need for a power supply59, and has been used frequently in arctic and peatland ecosystems19,25,60. The open-top chambers were fixed in place in early May of 2014 (in the center of each 2 m × 2 m plot), and the measurements ended in November 2015.
Air temperatures at vegetation canopy height were recorded continually at a 30-min time step using temperature loggers (Lascar Electronics, Salisbury, UK) and soil temperature at 5 cm and 20 cm depth was continuously recorded at a 30-min time step by soil temperature sensors (LI7900–180, LI-COR Inc., Lincoln, Nebraska, U.S.) connected to a Campbell data logger (CR1000, Campbell Scientific, Utah, USA) at two randomly chosen plots (one for the plot with warming and one for the plot with ambient temperature), and was also measured manually with a temperature probe in each plot during every gas sampling campaign. Water table levels were measured from dip-wells made of 1-m long perforated PVC pipe installed in each of the 64 experimental plots during every gas sampling campaign.
Annual background inorganic wet N deposition in the region is 0.5–0.6 g N m−261. For the N addition treatment, from the start of the study we annually applied 6.4 g N m−2, ~10 times the ambient annual wet N deposit. The rationale behind this quantity was to establish N-non-limited conditions for this nutrient-poor ecosystem, which is equivalent to the amount of high N addition level treatments used in another study in a northern peatland32. Nutrients were applied in soluble form as NH4NO3 in 2 L of water taken from an open pool close to the 2 × 2 m plots bimonthly from May to September each year. The same volume of open pool water was applied to the control sites.
Plant functional group manipulations were made only for the two dominant vascular vegetation types present at our sites: dwarf shrubs and graminoids. To avoid any soil disturbance, we did not manipulate the bryophyte/lichen functional type because of its substrate nature in this ecosystem. Vegetation removal was undertaken by hand from an area of 2 × 2 m. The shoots of shrubs and graminoids were cut back to litter layer level in early May 2014. Plots were left to settle for a year before sampling to minimize the effects of decomposition from roots. We did not include the first year’s data in our data analysis, since high variabilities of GHGs were observed for the first year (Supplementary Figure 3), and no significant treatment effect was statistically detected. Maintenance, such as removal of new shoots of removed vegetation types, was conducted regularly during each sampling campaign.
Measurements of gas exchange
In each sampling plot, a PVC (polyvinyl chloride) collar with an inner diameter of 26 cm was permanently inserted into the peat to a depth of 10 cm in the spring of 2014. The upper part of the collar features a groove to accommodate the water seal needed for the chamber measurements. Care was taken during insertion to minimize disturbance and to avoid severance of large plant roots. Boardwalks were installed to prevent any damage to the vegetation, disturbance to peat gas storage, or emissions during site visits. We had in total six sampling campaigns at ~3-week intervals from June to October 2015. During each sampling campaign, measurements were done for all the plots and conducted between 10:00–15:00 local time during 2–4 days to avoid rainfall.
For ecosystem respiration (Reco), CH4, and N2O fluxes, gas samples were collected using opaque chambers 50 cm in height and 26.3 cm in diameter, fitted to the groove of the PVC collar, covered with aluminum foil to reduce any solar heating effect, and equipped with a capillary tube to maintain atmospheric pressure inside the chamber when sampling. Samples were taken immediately upon closure of the chambers and at 10 min., 20 min. and 30 min. after closure. Flux was calculated by linear regression using all four measurements sampled during the 30 min39,57. Because the light was blocked by the opaque chamber, no photosynthesis occurred. Thus, the flux calculated based on the change of CO2 concentration inside the opaque chamber can be considered to be the ecosystem respiration. Gas samples (25 mL) were taken from the chamber headspace using a gas syringe and injected into pre-evacuated 12 mL Exetainer vials (Labco, Lampeter, UK) for storage prior to analysis. Concentrations of CO2, CH4, and N2O were analyzed by gas chromatography using a Scion 456-GC (gas chromatograph; Bruker, Milton, Canada) equipped with a thermal conductivity detector for CO2, a flame ionization detector for CH4, and an electron capture detector for N2O. For each sample, 5 mL of gas was injected into the chromatograph using an Autosampler (Combi PAL, Milton, Canada). The gas concentration was calculated using a calibration curve based on two certified standard gases, comprising 378 ppm and 0.303% CO2 (i.e., 3030 ppm), 2.52 and 17.7 ppm CH4, and 0.770 and 7.63 ppm N2O (Air Liquide, Canada). All fluxes were adjusted for field sampling temperature, headspace volume and chamber area62.
Soil pore water chemical composition
Soil pore water samples were collected using a MacroRhizons sampler (Rhizosphere, The Netherlands) installed at ~10 cm depth in each plot. The sampler has a porous tip with an outer diameter of 4.5 mm and a pore size of 0.15 µm. MacroRhizons have a female luer lock fitting, which is suitable for creating a vacuum with a male luer lock syringe63. Samples were collected during each gas sampling campaign. DOC and dissolved total nitrogen (TN) analysis was conducted on a Shimadzu TOC-LCPH/TN analyzer (Shimadzu, Japan). Three injections of each sample were applied to calculate the average DOC and TN concentration for each sample.
The composition of DOC was assessed using three indices: specific UV absorbance (SUVA254), defined as UV absorbance at 254 nm normalized for the DOC concentration (1 mg−1 C m−1), which increases linearly with measured DOC aromaticity 64. Reported values of SUVA254 in natural waters usually range from 0.5 to 6 l mg−1 C m−1, equivalent to a range of percent aromaticity between 5 and 45%;65 UV absorbance ratio, between 250 and 365 nm (a250/a365), which is inversely related to the DOC weight-averaged molecular weight66 with reported values in the range three to eight; reported C/N ratio, assumed equal to the measured DOC/TN ratio.
The effects of experimental warming, N addition, vegetation manipulation, and their interactions were analyzed by repeated measures ANOVA using IBM SPSS Statistics 20, with sampling date nested within sampling block as random effects. Vegetation manipulation effects were determined according to whether or not each of the two plant functional groups (shrubs or graminoids) was present. Data were checked for normality using the residual plots method and log-transformed where necessary before analysis.
Further information on experimental design is available in the Nature Research Reporting Summary linked to this article.
The data that support the findings of this study are available from the corresponding author upon reasonable request.
Yu, Z. Holocene carbon flux histories of the world’s peatlands Global carbon-cycle implications. Holocene 21, 761–774 (2011).
IPCC. Climate Change 2013: The Physical Science Basis. Contribution of Working Group I to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change. (Cambridge University Press, Cambridge, United Kingdom and New York, NY, USA 2013).
Moore, T. & Basiliko, N. In Boreal Peatland Ecosystems (eds. Wieder, R. K. & Vitt, D. H.) 125–143 (Springer Berlin Heidelberg, Berlin, Germany, 2006).
Frolking, S. & Roulet, N. T. Holocene radiative forcing impact of northern peatland carbon accumulation and methane emissions. Glob. Change Biol. 13, 1079–1088 (2007).
Fletcher, S. E. M., Tans, P. P., Bruhwiler, L. M., Miller, J. B. & Heimann, M. CH4 sources estimated from atmospheric observations of CH4 and its 13C/12C isotopic ratios: 1. Inverse modeling of source processes. Glob. Biogeochem. Cycles 18, GB4004 (2004).
Dise, N. B. Peatland response to global change. Science 326, 810–811 (2009).
Wu, J. & Roulet, N. T. Climate change reduces the capacity of northern peatlands to absorb the atmospheric carbon dioxide: the different responses of bogs and fens. Glob. Biogeochem. Cycles 28, 1005–1024 (2014).
Hobbie, S. E. Temperature and plant species control over litter decomposition in Alaskan tundra. Ecol. Monogr. 66, 503–522 (1996).
Hallinger, M., Manthey, M. & Wilmking, M. Establishing a missing link: warm summers and winter snow cover promote shrub expansion into alpine tundra in Scandinavia. New Phytol. 186, 890–899 (2010).
Tape, K., Sturm, M. & Racine, C. The evidence for shrub expansion in Northern Alaska and the Pan‐Arctic. Glob. Change Biol. 12, 686–702 (2006).
Elmendorf, S. C. et al. Plot-scale evidence of tundra vegetation change and links to recent summer warming. Nat. Clim. Change 2, 453 (2012).
Dieleman, C. M., Branfireun, B. A., McLaughlin, J. W. & Lindo, Z. Climate change drives a shift in peatland ecosystem plant community: Implications for ecosystem function and stability. Glob. Change Biol. 21, 388–395 (2015).
Elmendorf, S. C. et al. Global assessment of experimental climate warming on tundra vegetation: heterogeneity over space and time. Ecol. Lett. 15, 164–175 (2012).
Weltzin, J. F., Bridgham, S. D., Pastor, J., Chen, J. & Harth, C. Potential effects of warming and drying on peatland plant community composition. Glob. Change Biol. 9, 141–151 (2003).
Ward, S. E., Bardgett, R. D., McNamara, N. P., Adamson, J. K. & Ostle, N. J. Long-term consequences of grazing and burning on northern peatland carbon dynamics. Ecosystems 10, 1069–1083 (2007).
Strack, M., Waddington, J. M., Rochefort, L. & Tuittila, E. S. Response of vegetation and net ecosystem carbon dioxide exchange at different peatland microforms following water table drawdown. J. Geophys. Res. 111, G02006 (2006).
Updegraff, K., Bridgham, S. D., Pastor, J., Weishampel, P. & Harth, C. Response of CO2 and CH4 emissions from peatlands to warming and water table manipulation. Ecol. Appl. 11, 311–326 (2001).
Bridgham, S. D., Pastor, J., Dewey, B., Weltzin, J. F. & Updegraff, K. Rapid carbon response of peatlands to climate change. Ecology 89, 3041–3048 (2008).
Dorrepaal, E. et al. Carbon respiration from subsurface peat accelerated by climate warming in the subarctic. Nature 460, 616–619 (2009).
Ward, S. E., Bardgett, R. D., McNamara, N. P. & Ostle, N. J. Plant functional group identity influences short-term peatland ecosystem carbon flux: evidence from a plant removal experiment. Funct. Ecol. 23, 454–462 (2009).
De Deyn, G. B., Cornelissen, J. H. C. & Bardgett, R. D. Plant functional traits and soil carbon sequestration in contrasting biomes. Ecol. Lett. 11, 516–531 (2008).
Lange, M. et al. Plant diversity increases soil microbial activity and soil carbon storage. Nat. Commun. 6, 6707 (2015).
Petrie, M. D., Collins, S. L., Swann, A. M., Ford, P. L. & Litvak, M. E. Grassland to shrubland state transitions enhance carbon sequestration in the northern Chihuahuan Desert. Glob. Change Biol. 21, 1226–1235 (2015).
Turetsky, M. R. et al. A synthesis of methane emissions from 71 northern, temperate, and subtropical wetlands. Glob. Change Biol. 20, 2183–2197 (2014).
Ward, S. E. et al. Warming effects on greenhouse gas fluxes in peatlands are modulated by vegetation composition. Ecol. Lett. 16, 1285–1293 (2013).
Bellisario, L., Bubier, J., Moore, T. & Chanton, J. Controls on CH4 emissions from a northern peatland. Glob. Biogeochem. Cycles 13, 81–91 (1999).
Lai, D. Methane dynamics in northern peatlands: a review. Pedosphere 19, 409–421 (2009).
Bardgett, R. D., Manning, P., Morriën, E. & De Vries, F. T. Hierarchical responses of plant–soil interactions to climate change: consequences for the global carbon cycle. J. Ecol. 101, 334–343 (2013).
Hobbie, S. E. & Högberg, P. A synthesis: the role of nutrients as constraints on carbon balances in boreal and arctic regions. Plant Soil 242, 163–170 (2002).
Lund, M., Christensen, T. R., Mastepanov, M., Lindroth, A. & Ström, L. Effects of N and P fertilization on the greenhouse gas exchange in two northern peatlands with contrasting N deposition rates. Biogeosciences 6, 2135–2144 (2009).
Juutinen, S., Bubier, J. L. & Moore, T. M. Responses of vegetation and ecosystem CO2 exchange to 9 years of nutrient addition at Mer Bleue bog. Ecosystems 13, 874–887 (2010).
Larmola, T. et al. Vegetation feedbacks of nutrient addition lead to a weaker carbon sink in an ombrotrophic bog. Glob. Change Biol. 19, 3729–3739 (2013).
Heijmans, M. M. et al. Effects of elevated carbon dioxide and increased nitrogen deposition on bog vegetation in the Netherlands. J. Ecol. 89, 268–279 (2001).
Tomassen, H. B., Smolders, A. J., Limpens, J., Lamers, L. P. & Roelofs, J. G. Expansion of invasive species on ombrotrophic bogs: desiccation or high N deposition? J. Appl. Ecol. 41, 139–150 (2004).
Galloway, J. N. et al. Transformation of the nitrogen cycle: recent trends, questions, and potential solutions. Science 320, 889–892 (2008).
Fenner, N., Freeman, C. & Reynolds, B. Observations of a seasonally shifting thermal optimum in peatland carbon-cycling processes; implications for the global carbon cycle and soil enzyme methodologies. Soil Biol. Biochem. 37, 1814–1821 (2005).
Kalbitz, K., Schmerwitz, J., Schwesig, D. & Matzner, E. Biodegradation of soil-derived dissolved organic matter as related to its properties. Geoderma 113, 273–291 (2003).
Tveit, A. T., Urich, T., Frenzel, P. & Svenning, M. M. Metabolic and trophic interactions modulate methane production by Arctic peat microbiota in response to warming. Proc. Natl Acad. Sci. USA 112, E2507–E2516 (2015).
Luan, J. & Wu, J. Long-term agricultural drainage stimulates CH4 emissions from ditches through increased substrate availability in a boreal peatland. Agric. Ecosyst. Environ. 214, 68–77 (2015).
Liu, L. & Greaver, T. L. A review of nitrogen enrichment effects on three biogenic GHGs: the CO2 sink may be largely offset by stimulated N2O and CH4 emission. Ecol. Lett. 12, 1103–1117 (2009).
Saarnio, S., Saarinen, T., Vasander, H. & Silvola, J. A moderate increase in the annual CH4 efflux by raised CO2 or NH4NO3 supply in a boreal oligotrophic mire. Glob. Change Biol. 6, 137–144 (2000).
Silvola, J. et al. Effects of elevated CO2 and N deposition on CH4 emissions from European mires. Glob. biogeochem. cycles 17, 1068 (2003).
Banger, K., Tian, H. & Lu, C. Do nitrogen fertilizers stimulate or inhibit methane emissions from rice fields? Glob. Change Biol. 18, 3259–3267 (2012).
Hu, A. & Lu, Y. The differential effects of ammonium and nitrate on methanotrophs in rice field soil. Soil Biol. Biochem. 85, 31–38 (2015).
van Winden, J. F., Reichart, G.-J., McNamara, N. P., Benthien, A. & Damsté, J. S. S. Temperature-induced increase in methane release from peat bogs: a mesocosm experiment. PLoS One 7, e39614 (2012).
Curtis, C. J., Emmett, B. A., Reynolds, B. & Shilland, J. How important is N2O production in removing atmospherically deposited nitrogen from UK moorland catchments? Soil Biol. Biochem. 38, 2081–2091 (2006).
Jiang, G., Gutierrez, O., Sharma, K. R. & Yuan, Z. Effects of nitrite concentration and exposure time on sulfide and methane production in sewer systems. Water Res. 44, 4241–4251 (2010).
Roy, R. & Conrad, R. Effect of methanogenic precursors (acetate, hydrogen, propionate) on the suppression of methane production by nitrate in anoxic rice field soil. FEMS Microbiol. Ecol. 28, 49–61 (1999).
Bodelier, P. L., Roslev, P., Henckel, T. & Frenzel, P. Stimulation by ammonium-based fertilizers of methane oxidation in soil around rice roots. Nature 403, 421 (2000).
Schipper, L. A., Hobbs, J. K., Rutledge, S. & Arcus, V. L. Thermodynamic theory explains the temperature optima of soil microbial processes and high Q10 values at low temperatures. Glob. Change Biol. 20, 3578–3586 (2014).
Ettwig, K. F. et al. Nitrite-driven anaerobic methane oxidation by oxygenic bacteria. Nature 464, 543 (2010).
Dean, J. F. et al. Methane Feedbacks to the Global Climate System in a Warmer World. Rev. Geophys. https://doi.org/10.1002/2017RG000559 (2018).
Yvon-Durocher, G. et al. Methane fluxes show consistent temperature dependence across microbial to ecosystem scales. Nature 507, 488–491 (2014).
Olefeldt, D. et al. A decade of boreal rich fen greenhouse gas fluxes in response to natural and experimental water table variability. Glob. Change Biol. 23, 2428–2440 (2017).
Conant, R. T. et al. Sensitivity of organic matter decomposition to warming varies with its quality. Glob. Change Biol. 14, 1–10 (2008).
Jungkunst, H. F. & Fiedler, S. Latitudinal differentiated water table control of carbon dioxide, methane and nitrous oxide fluxes from hydromorphic soils: feedbacks to climate change. Glob. Change Biol. 13, 2668–2683 (2007).
Luan, J. & Wu, J. Gross photosynthesis explains the ‘artificial bias’ of methane fluxes by static chamber (opaque versus transparent) at the hummocks in a boreal peatland. Environ. Res. Lett. 9, 105005 (2014).
Marion, G. et al. Open‐top designs for manipulating field temperature in high‐latitude ecosystems. Glob. Change Biol. 3, 20–32 (1997).
Terrer, C., Vicca, S., Hungate, B. A., Phillips, R. P. & Prentice, I. C. Mycorrhizal association as a primary control of the CO2 fertilization effect. Science 353, 72–74 (2016).
Walker, M. D. et al. Plant community responses to experimental warming across the tundra biome. Proc. Natl Acad. Sci. USA 103, 1342–1346 (2006).
Reay, D. S., Dentener, F., Smith, P., Grace, J. & Feely, R. A. Global nitrogen deposition and carbon sinks. Nat. Geosci. 1, 430–437 (2008).
Holland, E. A. et al. In Standard soil methods for long-term ecological research (eds. Robertson, G. P., Coleman, D. C., Bledsoe, C. S. & Sollins, P.) 185–201 (Oxford University Press, Oxford, United Kingdom, 1999).
Strack, M. et al. Effect of water table drawdown on peatland dissolved organic carbon export and dynamics. Hydrol. Process. Int. J. 22, 3373–3385 (2008).
Weishaar, J. L. et al. Evaluation of specific ultraviolet absorbance as an indicator of the chemical composition and reactivity of dissolved organic carbon. Environ. Sci. & Technol. 37, 4702–4708 (2003).
Olefeldt, D. & Roulet, N. T. Effects of permafrost and hydrology on the composition and transport of dissolved organic carbon in a subarctic peatland complex. J. Geophys. Res. Biogeosci. 117, G01005 (2012).
Lou, T. & Xie, H. Photochemical alteration of the molecular weight of dissolved organic matter. Chemosphere 65, 2333–2342 (2006).
Special thanks are given to Prof. Henry Mann for his help in vegetation identification and Ms. Denise Bouzane from the Department of Natural Resources, NL, for her constructive guidance on the selection of our research site. We would also like to thank Mr. Jordan M. Chaffey and Mr. Mahammad Mumtaz for their assistance in field sampling and sample analysis in the lab. This study was made possible by the following sources of funding to J.W.: Natural Science and Engineering Research Council of Canada-Discovery Grant, Canada Foundation for Innovation-Leaders Opportunity Fund (CFI-LOF); the Ignite R&D, Leverage R&D, and Regional Collaboration Research Initiative (RCRI) of Research & Development Corporation, NL; Seed Bridge Fund of Memorial University; Agricultural Research Initiative of NL; Humber River Basin Research Initiative of NL; Grenfell Campus’ Start-up Research Fund and Vice-President Research Fund; and the following funding to J.L.: the National Natural Science Foundation of China (No. 31670450), the Fundamental Research Funds for ICBR (1632018004), the National Key R&D Program of China (2018YFC0507301; 2018YFD060010402). We would like to thank Dr. Thomas Halford for proofreading our manuscript.
The authors declare no competing interests.
Publisher’s note: Springer Nature remains neutral with regard to jurisdictional claims in published maps and institutional affiliations.
About this article
Cite this article
Luan, J., Wu, J., Liu, S. et al. Soil nitrogen determines greenhouse gas emissions from northern peatlands under concurrent warming and vegetation shifting. Commun Biol 2, 132 (2019). https://doi.org/10.1038/s42003-019-0370-1