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. 2023 Dec 8;167(4):523–543. doi: 10.1007/s10533-023-01103-1

Peatland restoration pathways to mitigate greenhouse gas emissions and retain peat carbon

Ülo Mander 1,✉, Mikk Espenberg 1, Lulie Melling 2, Ain Kull 1
PMCID: PMC11068583  PMID: 38707516

Abstract

Peatlands play a crucial role in the global carbon (C) cycle, making their restoration a key strategy for mitigating greenhouse gas (GHG) emissions and retaining C. This study analyses the most common restoration pathways employed in boreal and temperate peatlands, potentially applicable in tropical peat swamp forests. Our analysis focuses on the GHG emissions and C retention potential of the restoration measures. To assess the C stock change in restored (rewetted) peatlands and afforested peatlands with continuous drainage, we adopt a conceptual approach that considers short-term C capture (GHG exchange between the atmosphere and the peatland ecosystem) and long-term C sequestration in peat. The primary criterion of our conceptual model is the capacity of restoration measures to capture C and reduce GHG emissions. Our findings indicate that carbon dioxide (CO2) is the most influential part of long-term climate impact of restored peatlands, whereas moderate methane (CH4) emissions and low N2O fluxes are relatively unimportant. However, lateral losses of dissolved and particulate C in water can account up to a half of the total C stock change. Among the restored peatland types, Sphagnum paludiculture showed the highest CO2 capture, followed by shallow lakes and reed/grass paludiculture. Shallow lakeshore vegetation in restored peatlands can reduce CO2 emissions and sequester C but still emit CH4, particularly during the first 20 years after restoration. Our conceptual modelling approach reveals that over a 300-year period, under stable climate conditions, drained bog forests can lose up to 50% of initial C content. In managed (regularly harvested) and continuously drained peatland forests, C accumulation in biomass and litter input does not compensate C losses from peat. In contrast, rewetted unmanaged peatland forests are turning into a persistent C sink. The modelling results emphasized the importance of long-term C balance analysis which considers soil C accumulation, moving beyond the short-term C cycling between vegetation and the atmosphere.

Supplementary Information

The online version contains supplementary material available at 10.1007/s10533-023-01103-1.

Keywords: Carbon dioxide, Methane, Nitrous oxide, Paludiculture, Peatland restoration, Rewetting

Introduction

Importance of peatlands in a changing climate

Peatlands cover only about 3% of the Earth’s terrestrial surface (Gorham 1991) but play a crucial role in the global carbon (C) cycle. They act as significant C stores and are sources or sinks for greenhouse gases (GHG) like carbon dioxide (CO2), methane (CH4) and nitrous oxide (N2O; Frolking et al. 2011). These gases contribute to climate change and are considered crucial anthropogenic GHGs. Due to their high C density, peatlands are globally recognized as vital C reservoirs (Gallego-Sala et al. 2018) accounting for about 21% of the global soil organic C stock, estimated ~ 3000 Pg (Leifeld and Menichetti 2017). Peatlands also serve as substantial stores of organic N, with Northern peatlands alone accumulating 8–15 Pg N. When including tropical peatlands, the global estimations reached up to 26 Pg N (Swenson et al. 2019).

Undisturbed peatlands are currently a C sink (~ 0.1 Pg C y–1), a moderate source of CH4 (~ 0.03 Pg CH4 y–1), and a very weak source of N2O (~ 0.00002 Pg N2O–N y–1) (Frolking et al. 2011). However, anthropogenic activities, primarily agriculture and forestry drainage (up to 20% of global peatlands), result in net CO2 emissions (~ 0.1 Pg C y–1), reduced CH4 emissions (10% smaller than in natural conditions), and increased N2O emissions (~ 20 times higher than in natural peatlands). Consequently, subsidence and soil degradation contribute nearly 6% of global anthropogenic GHG emissions (Wichtmann et al. 2016). Most likely, the global peatland’s GHG balance has turned to a C source, a slightly smaller CH4 source, and a larger (but still small) N2O source (Frolking et al. 2011). In Europe, 46% of the remaining peatlands have degraded to the point where peat is no longer actively forming (Swenson et al. 2019).

As typical wetlands, peatlands are severely threatened by drainage, climate change, fires and groundwater extraction (Fluet-Chouinard et al. 2023). However, their restoration is beneficial, enabling C capture and sequestration, and minimizing N2O emissions (Leifeld and Menichetti 2017).

The northern hemisphere has experienced the highest warming during winter and early spring (Ljungqvist et al. 2016), leading to more frequent freeze–thaw cycles. Annual precipitation has also increased, particularly during the cold half-year in northern regions (Ljungqvist et al. 2016). Conversely, there are many indications of growing frequency of droughts during the warm season (Chiang et al. 2021). An increase in flash floods is also predicted (Zheng et al. 2022). These changing climatic conditions, droughts, rapid fluctuations in groundwater level (Mander et al. 2021), flash floods (Schindler et al. 2020), and soil moisture conditions (Pärn et al. 2018; Evans et al. 2021; Huang et al. 2021) can create hot spots and hot moments of GHG emission in peatland ecosystems.

Peatland restoration

Rewetting is a crucial step for conservation or sequestration of C in peatlands previously affected by drainage (Günther et al. 2020). Likewise, rewetting can affect nitrogen cycle and reduce cumulative N2O–N emissions by up to 70% in European peatlands (Liu et al. 2022). Usually, peatland restoration and rewetting are considered as synonyms, but in this paper we consider rewetting as increasing water table level compared with previously drained status. In case of restoration we expect that both water table and vegetation are manipulated in the way that enables to achieve ecosystem status similar to pre-drainage. Restoration pathways depend on factors such as initial vegetation of the drained area (Heger et al. 2022; Schaller et al. 2022), nutrient status of the residual peat layer (Kreyling et al. 2021) and the expected time frame for achieving planned ecological and socioeconomic benefits of the restoration.

Rewetting may transform former peat extraction sites into mires, paludicultural land, wet forests or shallow waterbodies. Ecologically, the preferred pathway is to restore them as mires (Wilson et al. 2022), which can be achieved by restoring water level, by a combination of rewetting and the application of peat moss layer transfer technique (Gonzalez-Sargas & Rochefort 2019), or to establish shallow waterbodies in hydrologically complex sites (Christen et al. 2016).

Peatland rewetting for agricultural use (including paludiculture) is challenging but may have a positive short-term effect on CO2 capture and GHG mitigation (Maljanen et al. 2010; Wilson et al. 2016a, b) but may not stop long-term peat mineralisation. Rewetting for paludiculture (Wichtmann et al. 2016) usually encompasses Sphagnum-based C farming (Gaudig et al. 2017/2018), energy crop cultivation (Hyvönen et al. 2009; Mander et al. 2012; Järveoja et al. 2016a; Kandel et al. 2020); wet forestry (Anadon-Rosell et al. 2022), or cultivation of the reed (Martens et al. 2021) or cattail (De Jong et al. 2021). Berries and other wetland plants are also suitable for paludiculture, but cranberry (Vaccinium) is the most suitable pioneer species for mire restoration and long-term C capture (Freeman et al. 2022).

Restoration of drained forests is usually achieved by blocking ditches and restoring water levels (Grand-Clement et al. 2015). Regulating the water regime in restored sites is effective for mitigation of GHG emissions (Järveoja et al. 2016b).

Peatland restoration is vulnerable to hydroclimatic conditions, particularly in the temperate zone. Changing precipitation pattern, increasing temperature and decreasing snowpack are expected to contribute to more frequent extreme events like droughts and torrential rains, resulting in increased vulnerability and interannual variability (Alm et al. 1999; Drollinger et al. 2019). In addition, it is also important to consider the potential substantial losses of dissolved and particulate C from drained and restored peatlands when estimating C budgets (Billett et al. 2010; Rosset et al. 2022).

Due to global warming, northern peatlands are projected to experience increased GHG emissions, particularly during non-growing period (Rafat et al. 2021), while Garcin et al. (2022) highlight a lack of knowledge of hydroclimatic vulnerability of peat C in tropical peatlands. Wet-dry seasonality of GHG fluxes is expected from tropical peatland soils (Inubushi et al. 2003). However, the overall impact of climate change on GHG fluxes needs to be better understood. Current understanding suggests that changes in soil temperature, photosynthesis, and soil moisture drive alterations in net C fluxes (Rafat et al. 2021).

Restoration versus afforestation of peatlands

Peatland restoration involves rewetting, whereas afforestation of drained peatlands and maintaining their drained condition cannot be considered equivalent to restoration. The key issue lies in the difference in the time scale and the discrepancy in distinguishing between short-term C capture in ecosystem (often observed in studies on GHG exchange between the atmosphere and ecosystem) and long-term C sequestration in soil. When peatlands are drained, C loss from the peat can offset the benefit of long-term CO2 sequestration achieved by afforestation (Jurasinski et al. 2020, 2023).

Currently, there is insufficient evidence on the long-term benefits of active afforestation of drained peatlands to mitigate climate change (Jurasinski et al. 2023). Afforestation on drained peatland forests and some former peat extraction areas may provide short-term benefits for climate change mitigation (Mäkiranta et al. 2007; Samariks et al. 2023). However, this approach does not account for the value of long-term C storage in peat. Similarly, intensive forestry on drained peatlands will not restore the peatland ecosystem's flora, fauna, and functions (Haapalehto et al. 2017; Loisel & Gallego-Sala 2022).

In addition, afforested drained peatlands are more susceptible to wildfires (Kohlenberg et al. 2018; Zheng et al. 2023). These risks are exacerbated by more frequent and more intense droughts in the boreal zone (Walker et al. 2019), resulting in losses of burnt wood and substantial C losses from burnt and burning peat layers (Liu et al. 2023). The impacts of severe fires have been devastating drained areas of formerly tropical peatland forests in Southeast Asia (Page et al. 2009).

Restoration-versus-afforestation of peatlands is being debated during the legislation procedure of the European Union Nature Restoration Law (NRL) (Jurasinski et al. 2023).

The general aim of this paper is to assess short-term C capture (GHG exchange between the atmosphere and the peatland ecosystem) and analyse the fluxes in the context of long-term C sequestration in peat of restored (rewetted) peatlands and afforested drained peatlands. As a specific objective we implement a conceptual model that compares changes in the long-term C budget and climate warming mitigation potential in the restored and afforested peatlands.

Conceptual framework

To provide a comprehensive understanding of the framework and to organize the rich source material, we have adopted a three-stage system. These stages refer to peatlands with different water regimes: natural (pristine), drained, and restored (rewetted) ecosystems. Within each stage of water regime, we further divided them into blocks based on primary land use types (Fig. 1). Estimated fluxes of all three GHGs–CO2, CH4 and N2O—are presented using a general three tier scale (high, medium and low fluxes). Flux values are represented by arrows, with upward-pointing arrows indicating emission and downward-pointing arrows indicating capture or uptake of the corresponding gas. In addition, we have indicated the averaged lateral losses of dissolved organic C in water (Fig. 1).

Fig. 1.

Fig. 1

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Conceptual figure showing main generalized types of peatland ecosystems from natural and drained to potentially restored peatlands. Proposed set-up for greenhouse gas emission from the main compartments of ecosystems is shown. Arrows indicate literature-based mean values of greenhouse gases CO2, CH4 and N2O fluxes (upward and downward arrows indicate emissions and captures of the corresponding gas, respectively). Question marks indicate scarce data without representative values. Aggregated lateral losses of dissolved organic carbon (DOC) in water are based on Wilson et al. 2016a. Rewetting is displayed as a necessary presumption for peatland restoration. Paludiculture—Sphagnum, berries and energy crops. Open restored areas – bog- and fen-type peatlands

Although our main focus lies on temperate and boreal peatlands, we also briefly discuss tropical peatland forests due to their significance as global hot spots of GHGs, and their large-scale escalating disturbance and destruction. In characterizing GHG fluxes in natural and drained peatlands, we primarily relied on literature sources supplemented by unpublished original data from our research group. However, for tropical wetlands, data is limited or sporadic, resulting in the use of question marks.

For restored (rewetted) peatlands, due to the scarcity and less systematic nature of the data, we present a selection of available GHG flux data in Table 1. The categories of restored (rewetted) peatlands are as in Fig. 1: wet forests, paludiculture, shallow lakes and open mires. Upon no accurate distinction between the mentioned categories in the source data we added a general category – rewetted peatlands – to both Table 1 and Fig. 2. The latter one presents average flux values (Fig. 2A–C) and GHG balances in CO2 equivalents (Fig. 2D and Fig. 2E). In the more general Fig. 1, paludiculture is shown as a single category. To compile this information, we referred to both literature sources and results from local research projects in Estonia. For our literature search, we utilized well-known indexing systems such as Web of Science (Clarivate Analytics), Scopus, and Google Scholar. We used following keywords and their combinations: “rewetted peatland(s)”, “restored wetland(s)”, “restored peatland(s)”, “rewetted forest(s)”, “paludiculture”, “Sphagnum farming”, “greenhouse gases, “carbon dioxide “, methane, “nitrous oxide”, “tropical peatlands”. In addition, we looked for references from cited rewetting-related papers.

Table 1.

Greenhouse gas fluxes in main types of boreal and temperate rewetted peatlands

Type Name Coordinates Country Period of study Nutrient status Plant community/ treatment NEE (CO2-C) CH4-C N2O-N Source
Rewetted peatlands
Twitchell Island old wetland (26 yrs), former agricultural land 38°06’N, 121°38’W USA, CA Aug. 2012–Aug. 2013 Rich

Schoenoplectus acutus, Typha latifolia, T. domingensis, T. angustifolia, Ludwigia

peploides, Lemna sp.)

 − 3970 387 Knox et al. 2015
Sherman Island young wetland, 13 yrs), former peatland pasture 38°05’N, 121°42’W USA, CA March 2012–Apr. 2013 Rich Schoenoplectus acutus, Typha spp., − 3680 530 Knox et al. 2015
Boreal zone aggregated Global review All rewetted peatlands (restored from drained forests, grssslands and croplands, and peat extraction sites) − 1300 123.6 0 Günther et al. 2020
Temperate zone aggregated Global review All rewetted peatlands (see above) − 400 205.9 0 Günther et al. 2020
Rewetted organic soils Germany Average of both nutrient-poor and nutrient-rich rewetted organic soils − 0.4 279 0.1 Tiemeyer et al. 2020
Rewetted peatlands Global review Average of 20 literature sources 675 131 1.2 Bianchi et al. 2021
Rewetted peatland forest
Rewetted boreal forests Global review Before 2016 Poor Coniferous and mixed peatland forests (soil fluxes) Selected examples from IPCC 2014 Wetlands Supplement − 1520 73 0.16 Wilson et al. 2016a
Rich Coniferous and mixed peatland forests (soil fluxes) − 1930 220 0.16 Wilson et al. 2016a
Rewetted temperate forests Global review Before 2016 Poor Coniferous, deciduous and mixed peatland forests (soil fluxes) − 1220 160 0.16 Wilson et al. 2016a
Rich 960 418 0.16 Wilson et al. 2016a
28 rewetted peatland forests Finland 1 year (2008–2019) Rich Fen forests (soil fluxes) 0.8 Minkkinen et al. 2020
Poor Bog forests (soil fluxes) 0.5 Minkkinen et al. 2020

Wöpkendorf

Alnus glutinosa

 54°7′36″N, 12°29′04″E Germany May 2018–Apr. 2020 Rich

Rewetted peatland alder forest;

Carex acutiformis, C. riparia, Hottonia palustres, Solanum dulcamara

 15 to 565 Köhn et al. 2021
Paludiculture
Reed/grasses

Lavassaare

Phalaris

 58°34′20″N, 24°23′ 15″E Estonia May 2010–May 2011 Rich Phalaris arundinacea fertilized 3.56 − 0,006 Mander et al. 2012
Rich P. arundinacea non-fertilized 1.75 − 0.03 Mander et al. 2012
Jan.-De.c 2014 Rich P. arundinacea fertilized 790 0.14 2.39 Järveoja et al. 2016a
Rich P.arundinacea non-fertilized 2010 0.18 1.08 Järveoja et al. 2016a

Keressaare

Phalaris

 58°36′56″N, 27°00′ 36″E Estonia Apr. 2015–Mar. 2018 Poor P.arundinacea high water table, fertilized, limed − 892 1.14 19.0 Maddison et al. 2018
P.arundinacea low water table, fertilized, limed − 249 3.07 28.1 Maddison et al. 2018

Gersloot

Typha latifolia

 53°01’N, 5°55’E Nether-lands Feb–June 2017 Rich Fen peat, lab experiments 72.9 2.18 Vroom et al. 2018
Øby P. arundinacea, Poa spp 56°27′32″N, 9°40′40″E Denmark Mar. 2015–Mar.2017 Rich P.arundinacea flooded plots − 3970 to − 5970 450 to 1,110 4.0 to 5.5 Kandel et al. 2020
Rich P.arundinacea semi-flooded plots − 2650 to − 4990 160 to 710 4.0 to 6.0 Kandel et al. 2020
Emergent crops Global review Average of 6 literature sources 174 1.68 Bianchi et al. 2021
Vaccinium spp. Saint-Louis-de-Blandford, 46°15’N, 72°00’W Canada, Quebec Growing season 2012, 2013 Poor Cultivated cranberry (Vaccinium macrocarpon) 27 0.61 Lloyd et al. 2019
Maima 58°35′54″N, 24°22′ 36″E Estonia (a) Poor Naturally regenerated cranberry (Oxycoccus palustris) − 799 68.5 2.0 Burdun et al. 2021
Laiuse 58°47′17″N, 26°31′ 47″E Estonia (a) Poor

Naturally regen. Cranberry

(Oxycoccus palustris)

 − 837 31.4 − 0.002 Viru et al. 2020; Burdun et al. 2021
Ess-soo 57°54′51″N, 26°41′ 26″E Estonia (a) Poor

Naturally regen. cranberry

(Oxycoccus palustris)

 − 707 5.4 0.08 Viru et al. 2020; Burdun et al. 2021
Sphagnum spp

Nordhümmlinger Moore

Rewetted former peat mining area

Sphagnum site

53

°02’ N, 7°29’

E

 Germany June 2010–Dec. 2011 Poor Sphagnum cuspidatum, Eriophorum angustifolium, Molinia caerulea − 949 233 − 2.58 Beyer & Höper 2015
Sphagnum cultivation June 2010–Dec. 2011 Poor S. papillosum, S. cuspidatum, S. palustre, S. fallax, E. angustifolium, Erica tetralix, Juncus effusus, Betula pendula, Drosera spp. − 987 24 0.55 Beyer & Höper 2015
Hankhauser Moor, rewetted former bog grassland 53°16’ N, 08°18’ E Germany Sep. 2011–ug. 2013 Poor Sphagnum palustre cultivation − 5470 to − 6290 10–14 0.01 Günther et al. 2017
Sep. 2011–Aug. 2013 Poor Sphagnum papillosum cultivation − 8750 to − 8980 12–27 − 1.0 to 1.0 Günther et al. 2017
Ess-soo 57°54′51″N, 26°41′ 26″E Estonia (a) Poor Dredged hollow with naturally regenerated Sphagnum − 702 380 0.07 Viru et al. 2020; Burdun et al. 2021
Maima 58°35′54″N, 24°22′ 36″E Estonia (a) Poor Dredged hollow with naturally regenerated Sphagnum − 740 116 − 0.11 Viru et al. 2020; Burdun et al. 2021
Provinzial-moor 52°40’ N, 07°06’ E Germany Mar. 2017–Mar. 2019 Rich

S. papillosum

S. palustre;

ditch irrigation

 − 600 to 2200 3 to 53 0.5 to 1.3 Oestmann et al. 2022
Drenth 52°41’ N, 07°05’ E Germany Rich

S. papillosum

drip irrigation

 700 to 900 1 to 4 1.9 to 12 Oestmann et al. 2022
Hankhauser Moor 53°16’ N, 08°18’ E Germany 2015–2022 Rich Sphagnum spp, production field − 1,609 ± 739 22.7 ± 6.7 − 0,47 ± 1.57 Daun et al. 2023
Shallow lake
Giel’cikau Kasyl 52°38′N, 25°21’E Belarus Aug. 2010- Aug. 2012 Rich Phragmites australis, Lemna spp − 4,453 to − 8,242 480 to 1470 0 to 0.005 Minke et al. 2016
Västkärr 59°06′ N, 14°45′ E Sweden 2010–2014, snow-free periods (65% yr) Rich Graminoids, Carex spp, Typha spp 44 4.3 Jordan et al. 2020

Zilakalna,

Tevgarsu

57°36’ N, 25°10’ E;

57°40’ N, 24°57’ E

 Latvia Dec. 2016–Dec. 2022 Poor

Permanently flooded peat extraction area;

no vegetation

 550 ± 5 251 ± 77 0.00 ± 0.06 Bardule et al. 2023
Laiuse 58°47′17″N, 26°31′ 47″E Estonia Sep. 2019–June 2023 Poor 14 ha, open water, P. australis, Typha & sedges in littoral 1,079 292 0.17 Burdun et al. 2021
Open bog/fen
Bois-des-Bel, restored bog 47°57’N, 69°26’W Canada, Quebec May–Oct. 2000, 2001, 2002 Poor

Sphagnum spp., Polytrichum spp., Chamaedaphne calyculata, Vaccinium angusti-folium, Ledum groenlandicum, E. vaginatum,

T. latifolia

 − 4,330 to − 10,253 Waddington et al. 2010
Wandering River restored bog (former horticultural area) 55°11’N, 112°29’W Canada, Alberta July–Sep. 2011, May–July 2012 Poor Sphagnum spp, Carex spp., E. vaginatum, Salix pedicellaris, Polytrichum strictum, Pohlia nutans − 11,647 to 1,095 − 3.23 to 540 Strack et al. 2014

Trebel River valley

Rewetted fen, former agric. grassland

Phragmites

 54°060’N, 12°44’E Germany March 2011–March 2012 Rich Phragmites australis − 830 to 320 110 Günther et al. 2015
Typha Rich Carex spp − 430 to 710 105 Günther et al. 2015
Carex Rich Typha spp. − 30 to 290 590–630 Günther et al. 2015

Himmel-moor

Rewetted in 1981

Heath site

 53°44′20’’N, 9°50′58’’E Germany Aug. 2010–Jan. 2012 Poor Erica tetralix, Calluna vulgaris, V. oxycoccus, Andromeda polifolia 84 478 − 0.16 Vanselow-Algan et al. 2015
Sphagnum Aug. 2010–Jan. 2012 Poor Sphagnum spp 16 748 0.34 Vanselow-Algan et al. 2015
Purple moor grass Aug. 2010–Jan. 2012 Poor Molinia caerulea 67 1.114 0.26 Vanselow-Algan et al. 2015

Burns Bog Ecol. Cons. Area (BBECA) Restored peat mining area

Rewetted Cleared

 49°06′37"N 123°00′03"W Canada, British Columbia June–Aug. 2014 Poor

Ledum groenlandicum, Betula pendula

V. corymbosum Sphagnum capillifolium,

 240 0.0075 Christen et al. 2016
Rewetted Sedge 49°07′09"N 123°00′01"W Canada, British Columbia June–Aug. 2014 Poor Rhynchospora alba, Dulichium arundinaceum. Sphagnum spp., L. groenlandicum, V. uliginosum 669 0.005 Christen et al. 2016
Bellacorick

54°07′ 30’’ N, 09

°33′ 22´´W

 Ireland Nov2008– Dec 2013 Poor

Sphagnum spp, Juncus effusus

Eriophorum spp.

 − 1,040 ± 800 90 ± 20 0 Wilson et al. 2016b
Tässi

58°32′ 16’’ N, 25

°51′ 43´´ E

 Estonia Mar. 2014–Mar. 2015 Poor Bryophytes, Sphagnum; high water level 50 2.0 ± 0.9 − 0.01 ± 0.02 Järveoja et al. 2016b
Poor Bryophytes, Sphagnum, herbs, shrubs; low water level − 247 0.9 ± 0.5 0.2 ± 0.1 Järveoja et al. 2016b
Seba Beach 53°27′ 17″ N, 114°52′50″W Canada, Alberta May–Sep. 2015 Poor All plots (moss, bare, Eriophorum) − 0.135 Brummell et al. 2017
Seba Beach 53° 33′N, 114° 44’W Canada, Alberta May–Aug. 2016, May–Aug. 2017 Poor

Sphagnum spp, sedges

(rest. 2009)

 21 Bieniada & Strack 2021
Poor Graminoids, Polytrichum moss, Sphagnum spp, sedges (rest. 2012) 703 Bieniada & Strack 2021
Bois-des-Bel 47°58′2″ N 69°25′43″ W Canada, Quebec Nov. 2013–Oct. 2014 Poor Sphagnum spp, sedges − 900 ± 180 44 ± 2 Nugent et al. 2018
Uchter Moor Rewetted since 1999 former peat mining area Germany Jan–Dec. 2017 Rich Sphagnum spp, Eriophorum vaginatum, Molinia caerulea E. angustifolium 262 ± 33 48.8 ± 9.8 0.57 ± 0.21 Schaller et al. 2022
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Negative values correspond to sequestration, positive values indicate emission to the atmosphere. Units: kg C ha−1 yr−1; kg N ha−1 yr−1

Knox et al.: g CO2–C m−2 yr−1; g CH4–C m−2 yr−1; eddy covariance (FMA analyser).

Günther et al. 2020: t CO2 ha−1 yr−1; kg CH4–C ha−1 yr−1; kg N2O–N ha−1 yr−1.

Bianchi et al.: t CO2-eq ha−1 yr−1; conversion factor of climate–carbon feedbacks: CH4 = 34 and N2O = 298 (Myhre et al. 2013).

Tiemeyer et al.: t CO2–C ha−1 yr−1; kg CH4–C ha−1 yr−1; kg N2O–N ha−1 yr−1, annual average.

Wilson et al. 2016a: t CO2-eq ha−1 yr−1; conversion factor of climate–carbon feedbacks: CH4 = 34 and N2O = 298 (Myhre et al. 2013).

Minkkinen et al.: g N2O–N m−2 yr−1; chambers, GC; average annual values.

Köhn et al.: kg CH4 ha−1 yr−1;chambers, GC, Los Gatos laser analyser.

Mander et al.: mg CO2–C m−2 h−1; μg CH4–C m−2 h−1; μg N2O–N m−2 h−1; chambers, GC, annual median values.

Järveoja et al. 2016a: g CO2–C m−2 h−1; g CH4–C m−2 h−1; chambers, GC, average annual values.

Maddison et al.: mg CO2–C m−2 h−1; µg CH4–C m−2 h−1; µg N2O–N m−2 h−1; chambers, GC; average annual values.

Vroom et al.: mg CH4 m−2 h−1; mg N2O m−2 d−1; chambers on mesocosms in lab, Picarro G2508.

Kandel et al.: Mg CO2–C ha−1 yr−1; Mg CH4–C ha−1 h−1; mg N2O–N m−2 h−1; chambers, GC; range between average annual values.

Lloyd et al.: μg CO2–C m−2 h−1; μg CH4–C m−2 h−1; μg N2O–N m−2 h−1; chambers, GC, vegetation period average values.

Burdun et al.: mg CO2–C m−2 h−1; µg CH4–C m−2 h−1; chambers, GC, average annual values.

Viru et al.: mg CO2–C m−2 h−1; µg CH4–C m−2 h−1; μg N2O–N m−2 h−1; chambers, GC, average monthly and annual values.

Oestmann et al. 2022: t CO2 ha−1 yr−1; g CH4–C m−2 yr−1; mg N2O–N m−2 h−1; chambers, GC, range between average annual values of sub-sites.

Beyer & Hõper 2015: g CO2 C m−2 yr−1; g CH4–C m−2 yr−1, mg N2O–N m−2 yr−1; chambers, GC, average annual values.

Günther et al. 2017: g CO2m−2 yr−1; g CH4 m−2 yr−1; mg N2O m−2 yr−1; chambers, GC, average annual.

Daun et al.: g CO2 m−2 yr−1; g CH4 m−2 yr−1; g N2O m−2 yr−1; chambers, GC, average annual ± sd.

Minke et al.: mg CO2–C m−2 h−1; µg CH4–C m−2 h−1; µg N2O–N m−2 h−1; chambers, LiC or (CO2); GC (CH4, N2O), average annual values.

Jordan et al.: mmol CH4 m−2 h−1; μmol N2O m−2 h−1; average annual values.

Bardule et al.: mg CO2–C m−2 h−1; mg CH4–C m−2 h−1; μg N2O–N m−2 h−1; chambers, GC, average annual ± se.

Waddington et al.: g CO2 m−2 d−1; chambers, IRGA analyser, growing season average values.

Strack et al.: g CO2 m−2 d−1; mg CH4 m−2 d−1chambers, IRGA analyser, growing season average values.

Günther et al.: g CO2 C m−2 yr−1; g CH4–C m−2 yr−1, chambers, GC, average annual values.

Vanselow-Algan et al.: g CO2 m−2 yr−1; μg CH4–C m−2 s−1; μg N2O–N m−2 s−1, chambers, GC, average annual values.

Christen et al.: g CO2m−2 d−1 with EC; nmol CH4 m−2 s−1 and nmol N2O–N m−2 s−1 with chambers & GC, June–August median values.

Wilson et al. 2016b: mg CO2–C m−2 h−1; mg CH4–C m−2 h−1; μg N2O–N m−2 h−1; chambers, GC, average annual ± sd.

Järveoja et al. 2016b: mg CO2–C m−2 h−1; μg CH4–C m−2 h−1; μg N2O–N m−2 h−1; chambers, GC, average annual values.

Brummell et al.: mg N2O-N m−2 d−1, chambers, GC, average annual values.

Bieniada & Strack: µg CH4–C m−2 h−1, Los Gatos; average annual values, chambers.

Nugent et al.: g CO2-C m−2 yr−1; eddy covariance; mean ± sd.

Schaller et al.: EC LiCor g CO2 m−2 yr−1; EC LosGatos g CH4 m−2 yr−1 and mg N2O-N m−2 yr−1, average annual ± standard error values.

aNEE CO2 measurements from July 2021 until June 2023: CO2 Reco, CH4 and N2O measurements from August 2017 until June 2023.

Fig. 2.

Fig. 2

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Average annual values of greenhouse gas fluxes in main groups of restored (rewetted) peatlands. A–C Annual average values of CO2-C (NEE), CH4-C and N2O fluxes. D: greenhouse gas balance in CO2equivalent GWP100 values (25 for CH4, 298 for N2O; Myhre et al. 2013). e: greenhouse gas balance in CO2equivalent GWP500 values (7.6 for CH4, 153 for N2O; Forster et al. 2007)

To explain the dilemma of peatland restoration and afforestation in a long-term perspective and mitigation of climate warming, we created a conceptual model that helps to characterize the changes in the C budget under different management practices (Fig. 3).

Fig. 3.

Fig. 3

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Conceptual figure of C balance in hemiboreal ombrotrophic afforested peatland sites in a 300-year time-span. For drained and restored peatlands, time since drainage is shown. The tabs on the managed peatland forests graphs represent 100-year logging cycles. See Supplementary Table 1 for input parameters and modelling assumptions

The resulting climatic effects of these management options are strongly time-dependent and rely on difference in forming above or below ground carbon pool and carbon turnover rate.

Our conceptual model of long-term C stock dynamics over 300-year period is based on the approach proposed by Minkkinen & Laine (1998). We developed it further to include C stock dynamics in nutrient-poor mires (open bog and bog forest), drained bog forest, and potential land use scenarios after milled peat extraction (restored bog, naturally regenerated unmanaged forest and managed forest). In modelling, we assumed equal C stock 686 t ha−1 in any peatland as initial status, corresponding to C stock with 2 m peat depth in pristine bog (Supplementary Table 1). As approximation, the stem, branches and coarse root biomass were assumed to have stable ratio (branches 12% of stem biomass and coarse roots 19% of above ground biomass) in any age group, wood density (kg m3) and C content (%) values were derived from literature (Kask and Pikk 2009; Külla 1997). Initial value of stem biomass was set 3 m3 ha− 1 for pristine open bog, 100 m3 ha−1 yr−1 for pristine bog forest and 0 m3 ha−1 for any other land use classes (initially harvested peatland forest or treeless peat extraction site) with annual stem biomass increment by 0 m3 ha−1 yr−1 for open bog, 0.1 m3 ha−1 yr−1 for stable continuous cover bog forest, 4 m3 ha−1 yr−1 for drained peatland forest and naturally regenerated forest on peat extraction site (initial value 0 m3 ha−1 yr−1 until year 20, thereafter gradually increasing until 4 m3 ha−1 yr−1 by year 50 and starting to decline to 0.1 m3 ha−1 yr−1 by year 160 as equal to pristine bog forest) while mean annual increment of 6 m3 ha−1 yr−1 was assigned to managed forest on drained former peat extraction site because of higher fertility of deeper peat layers (usually fen peat) after removal of Sphagnum peat. In pristine bog forest, restored bog and unmanaged peatland forest on former peat extraction site no harvest and timber removal is considered. Harvesting cycle is assumed to be 100 years in both restored managed peatland forest and drained managed peatland forests, while only stem biomass is removed. Fine root biomass was considered as part of C turnover and part of net ecosystem exchange (NEE). NEE values for modelled ecosystems are based on original data from Estonian study sites (Table 1), Estonian national GHG inventory, and literature (Salm et al. 2012; Wilson et al. 2016a). A conceptual scheme presents modelled long-term C stock dynamics in pristine mire ecosystems, restored and afforested drained peatlands (Fig. 3).

Results and discussion

Greenhouse gas fluxes in restored peatlands

Figure 1 illustrates a conceptual view of estimated GHG fluxes in natural, drained and restored (rewetted) peatlands. In restored peatlands, CO2 fluxes play the main role in climate impacts, while CH4 emissions are moderate and N2O fluxes are low making them less significant. Lateral losses of dissolved and particulate C in water can account up to a half of total C budget. Among the different restored peatlands, paludiculture has the highest estimated carbon dioxide (CO2) sequestration rate (> 1000 kg C ha−1 y−1) followed by rewetted forests, open fens and bogs. However, some rewetted peatlands are potential source of both CO2 and CH4, at least during the first 20 years after restoration (Vanselow-Algan et al. 2021).

According to our analysis, all restored peatlands were C sinks. However, there was significant variation of the data, mainly due to limited data availability and differences in the age since rewetting, as well as variation in plant cover development (Table 1). Average annual CO2–C stored in rewetted forests, open peatlands and paludicultural sites was − 928, − 534, and − 528 kg CO2–C ha−1 yr−1, respectively (Table 1). These findings are consistent with the CO2 flux values reported by Günther et al. (2020) for rewetted peatlands. Due to their diversity, paludicultural ecosystems showed the most significant variation in CO2-C fluxes ranging from high C capturing to moderate emissions, with the highest values observed in Phalaris arundinacea and Poa spp. plantation on rich fen peat (Kandel et al. 2020). Shallow lakes established on flooded peat extraction sites generally emitted low to moderate levels of CO2 except extensively vegetated eutrophic sites (Minke et al. 2016).

In all rewetted peatlands, the average annual CH4 fluxes were at moderate level, with the highest values observed in shallow lakes (451 kg CH4–C ha−1 yr−1) and followed by rewetted forests, open bogs/fens and paludiculture (218, 163, and 122 kg CH4–C ha−1 yr−1, respectively; Table 1). Average annual N2O fluxes ranged from low to moderate levels (0.01 to 4.45 kg N2O–N ha−1 yr−1), except in Phalaris paludiculture plantations on poor and acid peat where liming and fertilization resulted in fluxes of up to 19 kg N2O–N ha−1 yr−1 under high groundwater levels and 28.1 kg N2O–N ha−1 yr−1 under low water levels (Maddison et al. 2016; Table 1).

In management of drained peatlands, global warming potential (GWP) of GHG-s should be considered to avoid making decisions based solely on short-period benefits that may overlook the long-term climate cooling effect. Drained peatlands are known to be persistent CO2 emitters over the long term, while rewetted peatlands as resilient re-established mire ecosystems effectively contribute to mitigating climate change, even considering radiative forcing of increased CH4 emissions (Günther et al. 2020) and decreased N2O emissions. Figure 2D and E demonstrate the climate effect of rewetted peatland ecosystems in GWP100 and GWP500 timeframes. Mires are the only terrestrial ecosystems capable to continuously sequester atmospheric carbon in the long term (Gorham et al. 2012; Cobb et al. 2020), i.e., the carbon sequestration itself is more important than its compound (methane vs carbon dioxide).

For restored (rewetted) peatlands see literature sources in Table 1.

For natural reference and drained peatlands the following literature sources served as the basis of this figure: Abdalla et al. 2016; Aitova et al. 2023; Bardule et al. 2023; Bianchi et al. 2021; Bieniada & Strack 2021; Brummell et al. 2017; Burdun et al. 2021; Busman et al. 2023; Clement et al. 2020; Couwenberg et al. 2010; Daun et al. 2023; Frolking et al. 2011; Griffis et al. 2020; Günther et al. 2020; Hergoualc’h and Verschot 2014; Hyvönen et al. 2009; Inubushi et al. 2003; Järveoja et al. 2016a,b; Jauhiainen et al. 2008, 2012, 2019, 2023; Jordan et al. 2020; Kandel et al. 2020; Kull 2016; Maddison et al. 2016; Mander et al. 2008, 2012, 2016; Melling et al. 2007; Nugent et al. 2018; Oestmann et al. 2022; Oktarita et al. 2017; Pärn et al. 2023; Petrescu et al. 2015; Rosset et al. 2022; Sakata et al. 2015; Salm et al. 2012; Sjögestren et al. 2011; Takakai et al. 2006; Tang et al. 2018; Toma et al. 2011; Truu et al. 2020; Turetsky et al. 2014; Veber et al. 2021, 202X; Wilson et al. 2016a,b; Wong et al. 2018.

Modelling approach

To estimate the carbon balance of afforested drained and restored rewetted peatlands we developed a conceptual model. Based on the availability of data from our research projects and similar studies in other countries we chose hemiboreal ombrotrophic peatlands as a modelling example. Figure 3 presents estimated long-term dynamics of C stock of afforested drained and restored rewetted hemiboreal ombrotrophic peatlands and their reference pristine bog ecosystem. In a 300-year perspective, considering stable climate conditions, drained forest peat (former bogs) will lose 1.8 and 1.0 t C ha−1 yr−1, respectively, due to peat mineralization. In afforested peatlands under continuous drainage, C losses from peat reach 0.9 t C ha−1 yr−1, whereas ecosystem C losses (the peat + vegetation budget and 100-year timber-harvesting regime) are 0.4–0.6 t C ha−1 y−1. In comparison, fifty years after rewetting the naturally regenerated unmanaged peatland forest show decreased C loss, which is due to lower peat mineralization and C accumulation in biomass. In 160 years they achieve C dynamics similar to the pristine bog forest—with a moderate C sequestration rate of 0.1 t C ha−1 y−1. Restoring a peat extraction site to bog ecosystem would become C neutral in nearly 20 years, and onwards continuous mean annual C sequestration of 0.22 t C ha−1 y−1 is assumed (Supplementary Table 1).

Peatland forest drainage can give a significant increase in short-term C capture in biomass (Lohila et al. 2011), however, long-term dynamics in C stock and peat mineralisation remain largely unknown.

Restoration of crop plantations established on tropical peatlands and other dramatically altered peatlands differs from restoration of temperate and boreal bogs because the source community is predominantly swamp forest where peat-forming material is predominantly wood. Above ground and below ground litter formation under the reforestation and management with moderate drainage can more easily compensate peat mineralization than in temperate and boreal areas (Couwenberg et al. 2010).

Conclusions

Understanding the dynamics of GHG fluxes caused by land-use change is essential for successful peatland restoration. Our analysis identified several contradictory research results and gaps in a deep understanding of these processes. Notably, there is a lack of GHG flux data for most of tropical drained and restored peatlands, with the exception of oil palm plantations. Another important issue is retention of C in restored peatlands, where we can differentiate between short-term C capture (GHG exchange between the peatland and atmosphere) and long-term C capture (accumulation in soil). Based on our analysis of literature sources and own research results from Estonia, rewetting of former peat extraction areas for further management or conservation is the only viable approach for long-term C sequestration. In contrast, afforestation combined with continuous peatland drainage may have short-term economic benefits but leads to C losses in the long term.

Uncertainties in long-term estimations of C storage and GHG flux dynamics are remarkably high and do not allow for exact predictions. However, even educated guesses can be valuable for decision making on further management of peatlands. To make accurate estimations, it is crucial to investigate the full combined impact of hydroclimate change, microbial processes, and vegetation on GHG emissions from restored peatlands.

Supplementary Information

Below is the link to the electronic supplementary material.

Supplementary file1 (DOCX 16 kb) (16.9KB, docx)

Acknowledgements

The concept for this paper was developed at the workshop titled "Peatlands for climate change mitigation in agriculture" that took place in Aarhus, Denmark, on 4–5 October 2022, and which was sponsored by the Organisation for Economic Co-operation and Development (OECD) Co-operative Research Programme: Sustainable Agricultural and Food Systems. This study was supported by the European Research Council (ERC) under grant agreement No 101096403 (MLTOM23415R), the Estonian Research Council (PRG352, and MOBERC20 and MOBERC44), the European Commission through the European Regional Development Fund (Center of Excellence EcolChange, TK-131), and the EU programmes: HORIZON-CSA project No 101079192 “Living Labs for Wetland Forest Research (LiWeFor)”, LIFE21-IPC-EE-LIFE-SIP AdaptEst (MLTOM23090 (101069566) “Implementation of national climate change adaptation activities in Estonia”, and the Estonian State Forest Management Centre financed project LLTOM17250 "Water level restoration in cut-away peatlands: development of integrated monitoring methods and monitoring".

Author contributions

UM: conceptualized the study, acquired funding, analyzed the data, and drafted the manuscript. ME: conceptualized the study, analyzed the data, generated visualizations, and contributed to reviewing the manuscript. LM: contributed to reviewing the manuscript. AK: conceptualized the study, acquired funding, collected original data, analyzed the data, and contributed to reviewing the manuscript.

Funding

OECD Cooperative Research Programme,European Research Council,101096403,Ülo Mander, HORIZON EUROPE Widening Participation and Strengthening the European Research Area,101079192, Ülo Mander,Eesti Teadusagentuur,MOBERC20,Ülo Mander,MOBERC44,Ülo Mander,PRG352,Ülo Mander,European Regional Development Fund,EcolChange TK131,Ülo Mander,LIFE21-IPC-EE-LIFE-SIP, 101069566, Ain Kull,State Forest Management Centre,LLTOM17250, Ain Kull

Data availability

Data are available within the article and its supplementary materials. For additional data or questions, please contact the authors.

Footnotes

Publisher's Note

Springer Nature remains neutral with regard to jurisdictional claims in published maps and institutional affiliations.

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