# Long-Term Ecological Baselines and Critical Thresholds in Ombrotrophic Peatlands of Europe: Implications for Restoration Strategies

Mateusz Draga, Mariusz Galka, Klaus-holger Knorr, Stephan Glatzel, Bogdan H Chojnicki, Christian Fritz, Vincent E J Jassey, Radoslaw Juszczak, Hanna Meyer, Bjorn J M Robroek, et al.

► To cite this version:

Mateusz Draga, Mariusz Galka, Klaus-holger Knorr, Stephan Glatzel, Bogdan H Chojnicki, et al. Long-Term Ecological Baselines and Critical Thresholds in Ombrotrophic Peatlands of Europe: Implications for Restoration Strategies. Global Change Biology, 2025, 31 (12), pp.e70629. 10.1111/gcb.70629. hal-05412934

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RESEARCH ARTICLE

OPEN ACCESS

# Long-Term Ecological Baselines and Critical Thresholds in Ombrotrophic Peatlands of Europe: Implications for Restoration Strategies

Mateusz Draga<fnref n="1" />,² ID | Mariusz Gałka³ ID | Klaus-Holger Knorr⁴ ID | Stephan Glatzel⁵ ID | Bogdan H. Chojnicki⁶ | Christian Fritz⁷ | Vincent E. J. Jassey⁸ ID | Radosław Juszczak⁶ ID | Hanna Meyer⁹ | Bjorn J. M. Robroek⁷ ID | Carrie L. Thomas⁷ | Mariusz Lamentowicz<fnref n="1" /> ID

**Correspondence:** Mateusz Draga (matdra@amu.edu.pl) | Mariusz Lamentowicz (mariusz.lamentowicz@amu.edu.pl)

**Received:** 16 May 2025 | **Revised:** 3 October 2025 | **Accepted:** 29 October 2025

**Keywords:** change point | ecosystem threshold | peatland | resilience | *Sphagnum* | water table

## ABSTRACT

Maintaining appropriate peatland hydrology, notably through the regulation of the depth to water table (DWT), is crucial for peatland conservation, restoration, and the mitigation of greenhouse gas (GHG) emissions. We assess the long-term ecological impact of hydrological changes, primarily induced by drainage, in ombrotrophic peatlands across Europe. Our analysis is based on novel palaeoecological data from seven peat cores collected from sites that have experienced varying degrees of anthropogenic disturbance. We reconstructed historical DWT fluctuations using plant macrofossil and testate amoeba analyses at high resolution. By applying Threshold Indicator Taxa Analysis (TITAN) models, we identified species-specific and community-level response thresholds to changes in reconstructed water table. This approach revealed two distinct change points: the first, at c. 7 cm DWT, corresponds to hydrological conditions favourable for moisture-dependent *Sphagnum* species. The second, at c. 22 cm DWT, is associated with more drought-adapted plant taxa and signals ecosystem degradation. The interval between these change points represents a transition zone between optimal and suboptimal conditions for peatland functioning. An additional TITAN analysis, designed to identify the timing of major ecological changes, indicates that peatland degradation has intensified over the past two centuries and accelerated in recent decades. Our findings further reveal that plant and testate amoebae communities often remain distinct from those of undisturbed peatlands, even after hydrological restoration. This underscores the importance of preserving sites that still retain near-natural conditions. Based on our results (and consistent with previous studies) we recommend maintaining the water table close to the surface, i.e., a DWT of approximately 10 cm below the surface as an optimal target for both peatland conservation and restoration. Such

**Abbreviations:** DWT, depth to water table; GHG, greenhouse gases; TITAN, threshold indicator taxa analysis.

This is an open access article under the terms of the [Creative Commons Attribution License](https://creativecommons.org/licenses/by/4.0/), which permits use, distribution and reproduction in any medium, provided the original work is properly cited.

© 2025 The Author(s). *Global Change Biology* published by John Wiley & Sons Ltd.

<footnote n="1">¹Climate Change Ecology Research Unit, Faculty of Geographical and Geological Sciences, Adam Mickiewicz University in Poznań, Poznań, Poland | ²Department of Hydrobiology, Institute of Environmental Biology, Faculty of Biology, Adam Mickiewicz University Poznań, Poznań, Poland | ³Department of Biogeography, Paleoecology and Nature Conservation, Faculty of Biology and Environmental Protection, University of Lodz, Łodz, Poland | ⁴Institute of Landscape Ecology, Ecosystem Hydrology & Biogeochemistry Group, University of Münster, Münster, Germany | ⁵Department of Geography and Regional Research, Geoecology, University of Vienna, Vienna, Austria | ⁶Laboratory of Bioclimatology, Department of Ecology and Environmental Protection, Faculty of Environmental and Mechanical Engineering, Poznań University of Life Sciences, Poznan, Poland | ⁷Department of Ecology, Radboud Institute for Biological and Environmental Sciences, Faculty of Science, Radboud University Nijmegen, Nijmegen, the Netherlands | ⁸Université de Toulouse, Toulouse INP, CNRS, IRD, CRBE, Toulouse, France | ⁹Institute of Landscape Ecology, Remote Sensing & Spatial Modelling Group, University of Münster, Münster, Germany</footnote>
conditions not only support ecological integrity but are also associated with reduced GHG emissions and higher peat accumulation rates, reinforcing the role of ombrotrophic peatlands as long-term carbon sinks.

# 1 | Introduction

With ongoing climate change, peatlands remain one of our most valuable natural allies in our efforts to mitigate or offset carbon emissions (Strack et al. 2022). Although they occupy only 3% of Earth's land surface (Xu et al. 2018), peatlands support a wide range of essential ecosystem services and serve as important regulators of environmental processes (Kimmel and Mander 2010; Loisel et al. 2021). Intact peatlands contribute to groundwater stabilisation, enhance resilience to drought (Lennartz and Liu 2019; Taufik et al. 2020), exert localised cooling effects (Worrall et al. 2020), and help mitigate the risk of wildfires (Kettridge et al. 2017; Taufik et al. 2022). Moreover, peatlands often form biodiversity-rich microhabitats within otherwise intensively managed forest or agricultural landscapes. In addition to these ecological functions, peatlands play a pivotal role in the global carbon cycle. They are estimated to store between 30% and 44% of the world's soil carbon (Parish et al. 2008; Joosten et al. 2016), an amount comparable to the carbon currently present in the atmosphere (Yu 2012; Dargie et al. 2017). These vast carbon reservoirs have developed over tens of thousands of years through the slow accumulation of plant biomass. To slow the pace of global warming, it is imperative that this organic carbon remains stored in the peat (Strack et al. 2022). In this context, it is particularly alarming that many peatlands are now degraded, characterised by lowered water tables and a notable absence of active peat-forming processes (Page and Baird 2016; Swindles et al. 2019; Tanneberger, Abel, et al. 2021), due to decades, or even centuries of pressure from human activity on these ecosystems (Loisel et al. 2021). As a result of these changes, such sites not only lose their ability to absorb carbon from the atmosphere but, due to ongoing peat oxygenation, begin to emit significant amounts of greenhouse gases (GHG) back into the atmosphere (Leifeld et al. 2019; Doelman et al. 2023). Thus, considering both the benefits of well-preserved peatlands and the dangers posed by degraded ones, it is not surprising that increasing efforts are dedicated annually to various peatland protection and restoration projects (Apori et al. 2022). Regrettably, despite the existence of numerous restoration methods, such as drain blocking, rewetting, ditch blocking, tree harvesting, and in some cases re-introduction of target vegetation, herbicide spraying, and topsoil removal, peatland restoration remains a complex challenge (Artz et al. 2018; Zak and McInnes 2022; Stachowicz et al. 2023).

The functioning of peatlands largely depends on well-balanced and relatively stable hydrological conditions. Determining the depth to water table (DWT) (Baird and Low 2022) that supports both the functioning of intact peatlands and the regeneration of degraded ones is, however, a complex task (Lennartz and Liu 2019; Austin et al. 2025). That said, it should be noted that the DWT is not a fixed value and often varies significantly across seasons (Breeuwer et al. 2009). Despite these natural fluctuations, many peatlands have, in recent decades, experienced a sustained lowering of the average water table, leading to more frequent drought

conditions. This, in turn, has negatively affected the ecological niches of numerous species that are highly specialised and confined to peatland habitats. Over time, hydrophilic plant species typical of intact ecosystems may be replaced by species better adapted to lower or fluctuating water table and drier conditions, initiating a successional shift toward vascular plant dominance that ultimately accelerates peatland degradation and carbon loss (Jassey et al. 2018). Thus, ongoing plant species turnover on peatlands is usually a sign of rapidly progressing ecosystem degradation (Dieleman et al. 2015). Moreover, such turnover events do not solely affect plant communities, as such disturbances will also result in severe changes in invertebrate and microorganism species composition (Lamentowicz and Mitchell 2005; Swindles et al. 2016). However, because substantial shifts in community composition are typically well preserved in peat layers, these records offer researchers a valuable opportunity to reconstruct historical changes in the DWT and assess their long-term effects on species composition (Lamentowicz and Mitchell 2005; Chambers et al. 2012), thereby providing a stronger basis for informing ecosystem restoration efforts.

Linking changes in plant species composition from macrofossil data to proxies on concomitant changes in DWT may provide insight into the historical response of the plant community to drought (Booth et al. 2006; Swindles et al. 2010). The results can then be analysed to determine the critical DWT threshold beyond which peatland degradation begins (Bragazza 2008). Identifying such thresholds is crucial for effective peatland restoration and conservation efforts. While in most cases, predicting the value of a change point is a complicated task (Scheffer et al. 2009) if assessed through field studies (Jassey et al. 2018), fossil data provides us with a unique opportunity to estimate this parameter directly from the degradation events that happened in the past. Additionally, since paleodata can contain information spanning hundreds or even thousands of years, such data can provide researchers with hundreds of unique records on the relationship between plant community and peatland hydration to analyse. Moreover, these records are chronologically ordered, offering valuable insights into temporal changes and succession. Such an approach was utilised in the work of Lamentowicz et al. (2019), who estimated a plant community tipping point for the DWT value, analysing the fossil records collected from seven *Sphagnum*-dominated peatlands (raised bogs and poor fens) in Poland, with samples representing palaeoecological data from the past 2000 years.

To provide a geographically broader basis, the primary data for our analysis here are fossil records from seven ombrotrophic peatlands of diverse characteristics, spanning five European countries. Each site provides an average history spanning the past 2000 years. Notably, each site possesses a unique history, characterised by varying ages, preservation states, degradation levels, and exploitation histories. All of these sites have recently participated in restoration programs. This study aims to determine the critical transition zone for DWT in European ombrotrophic peatlands and identify the timing during which the most significant ecological change associated with drainage and peat
harvesting was observed. According to former palaeoecological and GHG observational studies (Lamentowicz et al. 2019; Evans et al. 2021; Fortuniak et al. 2021), we tested the hypothesis that the critical water table threshold is located at a depth of c. 10 cm and long-term shifts deviating from this value significantly affect vegetation. Going beyond former research, we focused on the disturbance time window and the context of the change point and ecological baselines related to the recent peatland land-use change.

## 2 | Methods

### 2.1 | Study Sites

The research is based on the data obtained from the analysis of seven different peat profiles, gathered from primarily *Sphagnum*-dominated peatlands located in Europe: Amtsvenn-Hündfelder Moor (Germany), Bagno Kusowo (Poland), Drebbersches Moor (Germany), Fochteloerveen (The Netherlands), Pichlmeier Moor (Austria), Pürgschachen Moor (Austria), and Store Mosse (Sweden) (Figure 1). The selected sites encompass peatlands with diverse histories and varying degrees of preservation. All

study sites were initially characterised by a wet, pristine phase preceding drainage. Furthermore, all analysed profiles show clear evidence of disturbance, visible both in the stratigraphic record and on the present-day surface.

#### 2.1.1 | Amtsvenn-Hündfelder Moor (Germany)

The Amtsvenn-Hündfelder Moor is a raised bog located in North Rhine-Westphalia, Germany, near the Dutch border, and has a long history of human impact. Significant manual peat extraction and agricultural reclamation began in the early 19th century, primarily for buckwheat cultivation. Despite these activities, the area was once renowned for its diverse lagg vegetation. However, in the 20th century, industrial peat extraction and, thereafter, large-scale reclamation efforts caused profound changes in the landscape. The construction of drainage ditches and peat cutting led to severe peat degradation and surface drying, resulting in vegetation shifts. Today, the site is protected as part of the Natura 2000 network. The core analysed in this study was extracted from an open *Sphagnum* lawn at which no harvesting had occurred, currently overgrown by *Erica tetralix*, *Calluna vulgaris*, and *Molinia caerulea*.

<figure><img src="image_3.png" /><figcaption>**FIGURE 1** | Locations of the seven sites from which peat cores were collected and subsequently analysed. FV—Fochteloërveen (Netherland), KU—Bagno Kusowo (Poland), DB—Drebbersches Moor (Germany), AV—Amtsvenn-Hündfelder Moor (Germany), PI—Pichlmeier Moor (Austria), PM—Pürgschachen Moor (Austria), SM—Store Mosse (Sweden). Map lines delineate study areas and do not necessarily depict accepted national boundaries.</figcaption></figure>
### 2.1.2 | Bagno Kusowo (Poland)

Bagno Kusowo is a raised bog located in the northwestern part of Poland and is recognised as one of the best examples of a Baltic-type raised bog in this country. Extensive peat extraction began in the late 19th century and continued into the early 20th century. By the early 1960s, the bog had been completely drained to facilitate further peat extraction, with most activity concentrated in the southern part of the peatland, which is now undergoing restoration. The northern section has remained waterlogged despite the drainage, likely preserving much of its original ecological characteristics. The site is protected as part of a nature reserve and the Natura 2000 network. The core analysed in this study was extracted from a former exploitation pond, currently overgrown by a diverse *Sphagnum* community.

### 2.1.3 | Drebbersches Moor (Germany)

Drebbersches Moor is a raised bog located in Lower Saxony in Germany that is a part of the larger raised bog area of Großes Moor. Before its allocation as a protected nature area, the site was used for buckwheat cultivation since the late 18th century and consequently subjected to surficial drainage and controlled burning. Dams have been built in the past decades to reduce water losses to adjacent peat harvesting areas surrounding the site. Yet, the site presumably still loses water to surrounding drainage structures within the peat harvesting area. In summer 2003, there was a fire at the site, which mainly affected the surface peat. The site is characterised by shallow water tables, reaching near-surface levels during the wettest times, yet on average still too dry for raised bog regeneration. The core analysed in this study was extracted from the last remaining near-natural part of the site in a moderately dry *Sphagnum papillosum* lawn.

### 2.1.4 | Fochteloërveen (Netherlands)

Fochteloërveen is a bog, located in the northern Netherlands, characterised by an accumulated peat layer reaching depths of up to 3 m. Historically, the site has been impacted by various disturbances, including fires, droughts, occasional flooding, and agricultural activities. Since 2000, it has been undergoing an intensive restoration process focused on large-scale rewetting efforts, which have resulted in partial moss regeneration. Despite these measures, Fochteloërveen's future remains threatened due to continued nutrient accumulation and recurring drought events. The entire area is designated as a national reserve and is protected under the Natura 2000 network. The core analysed in this study was extracted from a wet *Sphagnum* carpet in the peatland part, where restoration had been successful in establishing water tables close to the surface. The vegetation is dominated by *Sphagnum fallax*, *S. papillosum*, and *S. medium*, present under a rather dense cover of *Molinea caerulea*.

### 2.1.5 | Pichlmeier Moor (Austria)

Pichlmeier Moor is a bog located in the Austrian Enns valley in the eastern Alps, approximately 5 km east of Pürgschachen Bog. In the past, the area was extensively drained for peat extraction,

as evidenced by three heavily degraded vegetation ridges, between which peat was removed. The excavated sections between these ridges have been rewetted since the early 1990s, with some areas of open water in the centre between the ditches. Currently, the site is undergoing natural vegetation succession, forming floating *Sphagnum*-dominated peat mats in the extracted depressions. The core analysed in this study was collected from the edge of an exploitation pond ridge, thus capturing most of the intact peat profile prior to extraction, as well as the recently forming floating mat peat.

### 2.1.6 | Pürgschachen Moor (Austria)

Pürgschachen Moor is part of a bog complex in central Austria, situated in the Enns Valley of the eastern Alps, about 5 km west of Pichlmeier Moor. It is the largest relatively undisturbed raised valley bog in the Alps, covering an area of 62 ha. It is characterised by an intact treeless central area dominated by native peatland vegetation. However, the bog shows signs of degradation, as an advancing succession of *Betula pubescens* and *Pinus mugo* is observed. The core analysed in this study was extracted from an open *Sphagnum* lawn within the near-pristine part of the site.

### 2.1.7 | Store Mosse (Sweden)

The Store Mosse National Park is the largest relatively well-preserved peatland expanse in southern Sweden with several raised bog complexes embedded in fens and interspersed with open waters. Areas along a central ridge structure in this peatland complex were historically subjected to intensive peat extraction. The whole area is currently undergoing restoration, along with several other peatlands in Sweden, as part of the Addmire Project initiated in 2015. Store Mosse is characterised by well-preserved *Sphagnum* vegetation with a lawn-hollow structure. Declared a national park in 1982, the area is legally protected and included in the Natura 2000 network. The core was extracted from a wet *Sphagnum cuspidatum* lawn, with a sparse cover of *Rhynchospora alba* and *Andromeda polifolia*.

## 2.2 | Coring, Subsampling and Depth to Water Table Reconstruction

The cores from each site were sampled in 2022–2023 using a Wardenaar sampler (Wardenaar 1987) and an Instorf corer to recover a monolith from the upper peat layer. The samples were transported to the laboratory at the Adam Mickiewicz University, Poznań, Poland, and stored in a cold room (constant temperature of +4°C). Each core was subsampled for individual proxies, dated, and analysed at various resolutions. Pollen, microscopic charcoal, and testate amoebae were sampled at 5-cm intervals, whereas plant macrofossils and macroscopic charcoal were analysed continuously at 1-cm intervals.

The absolute chronology for each core was based upon 39 ¹⁴C AMS dates provided by the Poznań Radiocarbon Laboratory (Poland) (Table S1). The age-depth models were calculated using the OxCal 4.3 software (Bronk Ramsey 1995) applying the P-Sequences function with parameters: $k_0 = 0.6$, $\log_{10}(k/k_0) = 1$,
and interpolation = 1 cm (Bronk Ramsey 2008; Ramsey and Lee 2013). The IntCal20 atmospheric curve was used as the calibration dataset (Reimer et al. 2020).

High-resolution (1-cm peat slices of approximately 10 cm³ volume, in contiguous samples) plant macrofossil analysis was used to reconstruct local plant succession. The samples were washed and sieved under a warm water current over 0.20 mm mesh screens. The total of vegetative plant remains remaining on the sieve after rinsing the peat was set to 100%. The percentage of individual vascular plants and mosses was then estimated to the nearest 5%. The fossil carpological remains and vegetative fragments (leaves, rootlets, epidermis) were identified using identification keys (Smith 2004; Mauquoy and van Geel 2007) and compared to recently collected specimens. See Gałka et al. (2019) for more details on the methods used for plant macrofossil analysis of peat.

Peat material for testate amoeba community assessment (2 cm³ in volume) was sampled from the same depths as the microscopic charcoal analyses. Peat samples were washed under 0.3-mm sieves following the method described by Booth et al. (2010). Testate amoeba was analysed under a light microscope between 200× and 400× magnification, with a minimum of 100 tests per sample whenever possible (Payne and Mitchell 2008). Several keys and taxonomic monographs (Grospietsch 1958; Ogden and Hedley 1980; Meisterfeld 2001a, 2001b; Clarke 2003; Mazei and Tsyganov 2006) as well as internet resources (Siemensma 2019) were used to achieve the highest possible taxonomic resolution. Quantitative reconstruction of the testate amoeba-based DWT was performed in C2 software (Juggins 2003), using a transfer function approach applying the tolerance-downweighted weighted averaging model (Juggins and Birks 2012), and it was based on the European training set (Amesbury et al. 2016). Diagrams with palaeoecological proxy data were plotted using C2 software (Juggins 2003). DataGraph was used (MacAskill 2012) to draw the synthesis figure. As a result, the fossil testate amoeba-based quantitative reconstruction provides information on the average DWT in the vegetation season in the long-term context, depending on the peat accumulation rates.

## 2.3 | TITAN Threshold Analysis

Two threshold indicator taxa analyses (TITAN) were conducted to infer DWT and temporal ranges associated with plant species turnover and species-specific change points in response to DWT fluctuations (Baker and King 2010). TITAN is a statistical method that estimates both taxon-specific and community-level threshold values from ecological data, even for the data characterised by low occurrence for the number of different taxa—a common challenge in ecological studies (King and Baker 2014). TITAN classifies each taxon into two groups: taxa negatively responding (z−), and taxa positively responding (z+) to increases in the analysed environmental gradient (here, reconstructed DWT). Then, a community threshold is calculated separately for the z− and z+ groups based on the highest Sum(z) peak values along the environmental gradient. The accuracy of the results for each taxon is evaluated with purity and reliability scores. Purity indicates the proportion of times a taxon was assigned

to the final group, while reliability represents the proportion of times a taxon achieved a *p*-value <0.05 across bootstrap replicates. The final community threshold score can be calculated either from all analysed taxa (unfiltered data), regardless of their accuracy scores, or only from taxa that are both pure and reliable (filtered data). Although charcoal pieces are not taxa, they were also included in the analysis, as their presence may provide valuable insights into the potential relationship between extremely low DWT values and fire occurrences.

The first TITAN analysis assessed the influence of historical DWT levels on the presence of taxa recorded in the macrofossil data. For this purpose, data from all study sites were pooled and analysed collectively. Thresholds were calculated for all taxa that appeared at least three times in the entire dataset. The second TITAN analysis investigated the impact of time (years) on the presence of different taxa to identify potential community-level species turnover over the last few centuries. Data from six sites were used for this analysis, as the oldest layers from one site (Store Mosse) dated only to the mid-20th century. Additionally, since the analysis focused on the past few centuries, only samples from 1500 AD to the present were included. Furthermore, as in the previous analysis, only taxa occurring at least three times were included.

For all the calculations, the R program (R Core Team 2023) was used for which the R Studio environment (Posit team 2023) was utilised. During the data preparation and cleaning steps of the study, the *dplyr* package (Hadley et al. 2023) was used, while the threshold indicator taxa analysis was performed using the *Titan2* package (Baker et al. 2015).

# 3 | Results

## 3.1 | Age-Depth Models and Major Changes in Peat-Forming Plants

Radiocarbon dating indicated that the peat profiles varied considerably in age, ranging from just a few decades at Store Mosse to over 5000 years at Pichlmeier Moor, with the majority dating back at least 1500 years (Table 1, Table S1). The resulting datasets, which documented historical changes in DWT and plant species occurrences, were combined and used in subsequent analyses focusing on the recent disturbances related to drainage and peat harvesting. The full set of the age-depth models is presented in Figures S1–S7, complete plant macrofossil diagrams are presented in Figures S8–S14, and complete testate amoebae diagrams are presented in Figures S15–S21.

## 3.2 | Hydrology-Related Community Threshold Analysis for Peatland Taxa

The TITAN analysis of community thresholds for DWT values identified 11 taxa associated with wet conditions (z−) and 21 taxa associated with dry conditions (z+). The change points for individual taxa are presented in Table 2. Among the 32 taxa, 24 were identified as both pure and reliable indicators. The change points for unfiltered data were estimated at 7.4 cm DWT for taxa associated with wetter conditions (Sum(z−)
<table><caption>**TABLE 1** | Overview of all sites from which peat cores were extracted, including the country of origin, precise location of the peat core extraction site, brief usage history, state of preservation, age range of analysed material, and the number of analysed layers.</caption><thead><tr><th>Site name</th><th>Country</th><th>Location</th><th>History of use</th><th>State of preservation</th><th>Time range of a sample</th><th>No. of analysed samples</th></tr></thead><tbody><tr><td>Amtsvenn-Hündfelder Moor</td><td>Germany</td><td>52°10'31.3" N, 6°57'19.5" E</td><td>Drainage, mining</td><td>Poor/rewetting</td><td>-307 to 1982</td><td>89</td></tr><tr><td>Bagno Kusowo</td><td>Poland</td><td>53°48'45" N, 16°35'11.8" E</td><td>Drainage, mining</td><td>Moderate/rewetting</td><td>289-2020</td><td>99</td></tr><tr><td>Drebbersches Moor</td><td>Germany</td><td>52°41'59.8" N, 8°21'34.5" E</td><td>Drainage, mining</td><td>Poor/rewetting</td><td>637-2018</td><td>99</td></tr><tr><td>Fochteloer Veen</td><td>Netherlands</td><td>52°59'44.1" N, 6°23'39.3" E</td><td>Drainage, mining</td><td>Poor/rewetting</td><td>806-2020</td><td>92</td></tr><tr><td>Pichlmeier Moor</td><td>Austria</td><td>47°34'49.3" N, 14°24'55.3" E</td><td>Drainage, mining</td><td>Moderate/rewetting</td><td>-3370 to 2016</td><td>45</td></tr><tr><td>Pürgschachen Moor</td><td>Austria</td><td>47°34'50.0" N, 14°20'50.0" E</td><td>Drainage</td><td>Moderate</td><td>-339 to 2018</td><td>200</td></tr><tr><td>Store Mosse</td><td>Sweden</td><td>57°17'05.4" N, 14°00'03.9" E</td><td>Drainage</td><td>Good</td><td>1954-2018</td><td>44</td></tr></tbody></table>

species peak) and 22.1 cm DWT for taxa associated with drier conditions (Sum(z+) species peak) (Figures 2 and 3). Similarly, the change points derived from the filtered data were 5.8 cm DWT for Sum(z−) and 23.1 cm DWT for Sum(z+), showing only minor differences from the unfiltered data. The range between these values (Sum(z−) 7.4 and Sum(z+) 22.1) represents a potential critical transition zone between wet and dry equilibria, where peatland conditions shift from preserved ecosystems to degraded ones. Among the taxa recognised as both pure and reliable, *Sphagnum papillosum*, *S. cuspidatum*, *S. austinii*, *S. medium/divinum*, *S. angustifolium*, *Andromeda polifolia*, *Rhynchospora alba*, and *Scheuchzeria palustris*, were assigned to the z− taxa group and taxa such as *S. russowii*, *S. fuscum/rubellum*, *S. palustre*, *Calluna vulgaris*, *Ericaceae*, fungal remains of sclerotia, *Erica tetralix*, *Poaceae*, *Hylocomnium splendens*, *Oxycoccus* sp., shrub rootlets, *Empetrum* sp., *Pleurozium schreberi*, *mycorrhizal roots*, *Drosera rotundifolia* and charcoal pieces to the z+ group (taxa in both groups are listed in order of increasing Zenv.cp values; Table 2).

### 3.3 | Time-Related Community Threshold Analysis for Peatland Taxa

The TITAN threshold analysis for the presence of peatland taxa over time, spanning from the year 1500 AD to the present, identified 15 taxa associated with older periods (z−) and 11 taxa linked to more modern timeframes (z+). Among these, seven taxa from the z− group and five from the z+ group were classified as both pure and reliable. The change points for individual taxa are provided in Table 3. The community change points for each group in the unfiltered dataset were identified at 1793 AD for Sum(z−) and 2014 AD for Sum(z+), respectively (Figures 4 and 5), and at 1793 AD and 2013 AD for the filtered dataset. The interval between these values represents a transition zone between historically dominant plant taxa communities and those currently

observed. Distinct peaks for both Sum(z−) and Sum(z+) underscore the statistical significance of these thresholds. Among the taxa identified as both pure and reliable, *Trichophorum/Eriophorum*, *Warnstorfia fluitans*, *Scheuchzeria palustris*, *Sphagnum cuspidatum*, *Rhynchospora alba*, *S. medium/divinum*, and charcoal pieces were placed within the z− group and the following taxa in the z+ group: *Poaceae*, *S. papillosum*, *Erica tetralix*, *Ericaceae*, *Oxycoccus* sp., and *S. palustre* (taxa in both groups are listed in order of increasing Zenv.cp values) (Table 3).

### 3.4 | Effect of Changing DWT on the Plant Taxonomic Composition in Peat Cores

The analysis of plant macrofossil composition found in peat cores, combined with testate amoebae-derived DWT values, revealed significant differences in the hydrological histories of the seven studied peatlands (Figure 6). All sites experienced substantial fluctuations in DWT, with some changes particularly prominent. Among them, only Bagno Kusowo exhibited an improvement in its DWT value in the second half of the 20th century, with DWT decreasing from approximately 20 cm to around 12 cm below the surface, presumably an effect of the exploitation pond overgrowing by different plant communities. This shift was accompanied by a transition in the dominant *Sphagnum* species, from *S. fuscum/rubellum* (classified in the first TITAN analysis as a drought-adapted taxon from the z+ group, Table 2) to *S. medium/divinum* (associated with the more water-dependent z− group).

An especially notable and abrupt decline in the water table (increase in DWT), leading to significant shifts in dominant plant taxa, was observed at Amtsvenn-Hündfelder Moor and Drebbersches Moor. In these sites, the DWT values changed from around 5 to 25 cm and from −5 (above surface) to 25 cm, respectively. At Amtsvenn-Hündfelder Moor, *S. austinii*
<table><caption>TABLE 2 | TITAN-calculated DWT change points for both z- and z+ taxa.</caption><thead><tr><th>Name of the taxon</th><th>Zenv.cp</th><th>Ienv.cp</th><th>Z score</th><th>IndVal</th><th>Frequency</th><th>Purity</th><th>Reliability</th></tr></thead><tbody><tr><td colspan="8">z- taxons</td></tr><tr><td><strong>Sphagnum papillosum</strong></td><td>-2.42</td><td>-2.51</td><td>21.88</td><td>36.14</td><td>81</td><td>1.000</td><td>1.000</td></tr><tr><td><strong>Sphagnum cuspidatum</strong></td><td>-2.12</td><td>-7.81</td><td>28.30</td><td>57.11</td><td>160</td><td>1.000</td><td>1.000</td></tr><tr><td><strong>Andromeda polifolia</strong></td><td>-1.44</td><td>-6.83</td><td>8.71</td><td>11.19</td><td>40</td><td>0.976</td><td>0.998</td></tr><tr><td><strong>Sphagnum angustifolium</strong></td><td>3.62</td><td>3.52</td><td>11.17</td><td>11.39</td><td>36</td><td>1.000</td><td>1.000</td></tr><tr><td><em>Warnstorfia fluitans</em></td><td>5.18</td><td>5.15</td><td>4.58</td><td>2.46</td><td>7</td><td>1.000</td><td>0.926</td></tr><tr><td><strong>Rhynchospora alba</strong></td><td>5.25</td><td>-3.43</td><td>5.86</td><td>6.73</td><td>35</td><td>1.000</td><td>1.000</td></tr><tr><td><em>Carex limosa</em></td><td>7.07</td><td>6.87</td><td>2.02</td><td>0.95</td><td>3</td><td>0.928</td><td>0.400</td></tr><tr><td><strong>Scheuchzeria palustris</strong></td><td>7.23</td><td>6.87</td><td>10.47</td><td>9.37</td><td>31</td><td>1.000</td><td>1.000</td></tr><tr><td><em>Pinus mugo</em></td><td>12.51</td><td>16.37</td><td>2.32</td><td>2.04</td><td>10</td><td>0.868</td><td>0.780</td></tr><tr><td><strong>Sphagnum austinii</strong></td><td>12.80</td><td>12.51</td><td>8.73</td><td>14.50</td><td>76</td><td>1.000</td><td>1.000</td></tr><tr><td><strong>Sphagnum medium/ divinum</strong></td><td>17.50</td><td>25.32</td><td>17.14</td><td>58.65</td><td>433</td><td>1.000</td><td>1.000</td></tr><tr><td colspan="8">z+ taxons</td></tr><tr><td><em>Trichophorum/Eriophorum</em></td><td>-6.49</td><td>-35.16</td><td>3.43</td><td>48.29</td><td>311</td><td>0.836</td><td>0.990</td></tr><tr><td><strong>Sphagnum russowii</strong></td><td>2.25</td><td>19.08</td><td>7.87</td><td>10.08</td><td>47</td><td>1.000</td><td>1.000</td></tr><tr><td><strong>Sphagnum fuscum/ rubellum</strong></td><td>6.24</td><td>35.01</td><td>20.85</td><td>43.30</td><td>217</td><td>1.000</td><td>1.000</td></tr><tr><td><strong>Calluna vulgaris</strong></td><td>6.82</td><td>-7.61</td><td>8.47</td><td>8.90</td><td>37</td><td>0.976</td><td>1.000</td></tr><tr><td><strong>Ericaceae</strong></td><td>8.29</td><td>6.19</td><td>19.34</td><td>26.59</td><td>99</td><td>1.000</td><td>1.000</td></tr><tr><td><em>Drepanocladus sp.</em></td><td>12.49</td><td>23.23</td><td>4.62</td><td>2.16</td><td>6</td><td>0.994</td><td>0.940</td></tr><tr><td><strong>Fungal remains</strong></td><td>15.03</td><td>15.18</td><td>7.15</td><td>5.34</td><td>17</td><td>1.000</td><td>1.000</td></tr><tr><td><strong>Erica tetralix</strong></td><td>17.62</td><td>-7.75</td><td>9.78</td><td>7.74</td><td>20</td><td>0.972</td><td>1.000</td></tr><tr><td>Herbs</td><td>20.48</td><td>-7.75</td><td>8.44</td><td>32.87</td><td>208</td><td>0.488</td><td>1.000</td></tr><tr><td><strong>Sphagnum palustre</strong></td><td>22.42</td><td>27.79</td><td>25.09</td><td>26.59</td><td>47</td><td>1.000</td><td>1.000</td></tr><tr><td><strong>Poaceae</strong></td><td>22.62</td><td>22.86</td><td>12.37</td><td>10.17</td><td>20</td><td>1.000</td><td>1.000</td></tr><tr><td><strong>Hylocomnium splendens</strong></td><td>24.92</td><td>29.49</td><td>12.34</td><td>7.81</td><td>10</td><td>1.000</td><td>1.000</td></tr><tr><td><em>Oxycoccus sp.</em></td><td>25.06</td><td>27.79</td><td>10.30</td><td>15.54</td><td>53</td><td>0.998</td><td>1.000</td></tr><tr><td><strong>Shrubs rootlets</strong></td><td>25.32</td><td>26.70</td><td>10.48</td><td>54.68</td><td>371</td><td>1.000</td><td>1.000</td></tr><tr><td><em>Empetrum sp.</em></td><td>26.74</td><td>26.82</td><td>12.01</td><td>6.86</td><td>7</td><td>1.000</td><td>0.994</td></tr><tr><td><strong>Pleurozium schreberi</strong></td><td>29.10</td><td>29.76</td><td>13.20</td><td>9.56</td><td>8</td><td>1.000</td><td>0.998</td></tr><tr><td><em>Pinus sylvestris</em></td><td>29.10</td><td>-7.82</td><td>6.63</td><td>10.81</td><td>25</td><td>0.884</td><td>0.990</td></tr><tr><td><strong>Mycorrhizal roots</strong></td><td>29.58</td><td>29.76</td><td>4.87</td><td>6.44</td><td>12</td><td>0.990</td><td>0.956</td></tr><tr><td><strong>Drosera rotundifolia</strong></td><td>30.54</td><td>31.45</td><td>11.25</td><td>12.19</td><td>6</td><td>1.000</td><td>1.000</td></tr><tr><td><strong>Charcoal pieces</strong></td><td>32.92</td><td>35.09</td><td>9.93</td><td>22.56</td><td>23</td><td>0.972</td><td>1.000</td></tr><tr><td><em>Sphagnum sect. Acutifolia</em></td><td>34.19</td><td>34.19</td><td>4.84</td><td>12.21</td><td>10</td><td>0.990</td><td>0.928</td></tr></tbody></table>

Note: Bold text highlights taxa that are both pure and reliable.

Abbreviations: Frequency, number of subsamples with the species present; Ienv.cp, DWT change point based on the indicator value score; IndVal, indicator value score for an individual species; Purity, proportion of bootstrap replicates assigning a taxon to the same category; Reliability, proportion of bootstrap replicates with a p-value <0.05 for a given taxon, indicating a predictable response to the DWT gradient; Z score, Z score value for an individual species; Zenv.cp, DWT change point based on the Z score.
<figure><img src="image_4.png" /><figcaption>**FIGURE 2** | Plot showing taxon-specific change points along the DWT gradient for taxa associated with wet (z−) and dry (z+) conditions. Positive values of the DWT parameter indicate a water table below the surface, while negative values represent water accumulation above ground level.</figcaption></figure>

<figure><img src="image_5.png" /><figcaption>**FIGURE 3** | Plot showing TITAN sum(z−) and sum(z+) values for all potential DWT change points. Positive values of the DWT parameter indicate a water table below the surface, while negative values represent water accumulation above ground level. Peaks in sum(z) indicate the point beyond which synchronous declines in the summarised z-score values occur within the given taxon group. Dot size is proportional to the Z score of a given taxon.</figcaption></figure>

(z− group), which thrived under more favourable hydrological conditions, disappeared, being replaced by *S. fuscum/rubellum* (z+ group) along with an increased presence of shrub rootlets (z+). A similar pattern was recorded at Drebbersches Moor, where an increase in DWT led to the replacement of *S. cuspida-tum* and *S. medium/divinum* (both z− taxa) by *S. papillosum* and the emergence of vascular plants such as *Erica tetralix* (z+), as

well as herbs (z+) and additional shrubs (z+), particularly over the last few decades.

Despite covering only the last few decades, the core of Store Mosse also revealed an intriguing shift in dominant *Sphagnum* species. For several years, *S. cuspidatum* (z−) was the dominant species in the analysed profile. However, following an increase in DWT
<table><caption>TABLE 3 | TITAN-calculated time change points for both z- and z+ taxa.</caption><thead><tr><th>Name of the taxon</th><th>Zenv.cp</th><th>Ienv.cp</th><th>Z score</th><th>IndVal</th><th>Frequency</th><th>Purity</th><th>Reliability</th></tr></thead><tbody><tr><td colspan="8">z- taxons</td></tr><tr><td><strong>Trichophorum/Eriophorum</strong></td><td><strong>1660.85</strong></td><td><strong>1646.2</strong></td><td><strong>8.12</strong></td><td><strong>69.64</strong></td><td><strong>67</strong></td><td><strong>1.000</strong></td><td><strong>0.998</strong></td></tr><tr><td>Carex limosa</td><td>1690.35</td><td>1690.35</td><td>9.14</td><td>13.04</td><td>3</td><td>0.946</td><td>0.784</td></tr><tr><td>Andromeda polifolia</td><td>1736.95</td><td>1610.50</td><td>4.93</td><td>11.04</td><td>6</td><td>0.906</td><td>0.910</td></tr><tr><td><strong>Warnstorfia fluitans</strong></td><td><strong>1772.45</strong></td><td><strong>1736.95</strong></td><td><strong>10.24</strong></td><td><strong>17.95</strong></td><td><strong>7</strong></td><td><strong>1.000</strong></td><td><strong>0.998</strong></td></tr><tr><td><strong>Scheuchzeria palustris</strong></td><td><strong>1787.50</strong></td><td><strong>1793.70</strong></td><td><strong>23.27</strong></td><td><strong>71.28</strong></td><td><strong>31</strong></td><td><strong>1.000</strong></td><td><strong>1.000</strong></td></tr><tr><td><strong>Sphagnum cuspidatum</strong></td><td><strong>1793.70</strong></td><td><strong>1781.30</strong></td><td><strong>21.04</strong></td><td><strong>60.05</strong></td><td><strong>30</strong></td><td><strong>1.000</strong></td><td><strong>1.000</strong></td></tr><tr><td><strong>Rhynchospora alba</strong></td><td><strong>1799.90</strong></td><td><strong>1574.75</strong></td><td><strong>10.56</strong></td><td><strong>27.21</strong></td><td><strong>17</strong></td><td><strong>1.000</strong></td><td><strong>1.000</strong></td></tr><tr><td><strong>Sphagnum medium/divinum</strong></td><td><strong>1954.15</strong></td><td><strong>1616.75</strong></td><td><strong>4.42</strong></td><td><strong>42.46</strong></td><td><strong>116</strong></td><td><strong>0.976</strong></td><td><strong>0.994</strong></td></tr><tr><td><strong>Charcoal pieces</strong></td><td><strong>1957.45</strong></td><td><strong>1957.45</strong></td><td><strong>5.40</strong></td><td><strong>9.41</strong></td><td><strong>8</strong></td><td><strong>1.000</strong></td><td><strong>0.982</strong></td></tr><tr><td>Drepanocladus sp.</td><td>1964.25</td><td>1963.40</td><td>3.63</td><td>6.38</td><td>6</td><td>0.998</td><td>0.832</td></tr><tr><td>Shrubs rootlets</td><td>1965.50</td><td>1605.25</td><td>4.50</td><td>48.34</td><td>115</td><td>0.840</td><td>1000</td></tr><tr><td>Hylocomnium splendens</td><td>1983.20</td><td>1982.70</td><td>2.10</td><td>5.04</td><td>6</td><td>0.914</td><td>0.604</td></tr><tr><td>Pleurozium schreberi</td><td>1992.70</td><td>1992.7</td><td>2.09</td><td>6.06</td><td>8</td><td>0.844</td><td>0.628</td></tr><tr><td>Herbs</td><td>1993.30</td><td>2016.10</td><td>4.31</td><td>34.63</td><td>73</td><td>0.786</td><td>0.998</td></tr><tr><td>Empetrum sp.</td><td>1998.30</td><td>1998.30</td><td>1.35</td><td>4.90</td><td>7</td><td>0.730</td><td>0.450</td></tr><tr><td colspan="8">z+ taxons</td></tr><tr><td><strong>Sphagnum fuscum/rubellum</strong></td><td>1743.30</td><td>2019.15</td><td>3.97</td><td>46.30</td><td>100</td><td>0.796</td><td>1000</td></tr><tr><td>Mycorrhizal roots</td><td>1,843.359</td><td>1852.90</td><td>1.85</td><td>6.12</td><td>9</td><td>0.744</td><td>0.594</td></tr><tr><td>Calluna vulgaris</td><td>1909.75</td><td>1697.70</td><td>2.90</td><td>14.27</td><td>27</td><td>0.962</td><td>0.918</td></tr><tr><td><strong>Poaceae</strong></td><td><strong>1958.10</strong></td><td><strong>2014.75</strong></td><td><strong>4.30</strong></td><td><strong>13.49</strong></td><td><strong>20</strong></td><td><strong>1.000</strong></td><td><strong>0.982</strong></td></tr><tr><td><strong>Sphagnum papillosum</strong></td><td><strong>1958.55</strong></td><td><strong>1958.55</strong></td><td><strong>5.24</strong></td><td><strong>13.84</strong></td><td><strong>17</strong></td><td><strong>0.988</strong></td><td><strong>1.000</strong></td></tr><tr><td>Erica tetralix</td><td>1959.15</td><td>1959.15</td><td>3.37</td><td>12.08</td><td>19</td><td>0.962</td><td>0.932</td></tr><tr><td>Pinus sylvestris</td><td>1976.30</td><td>2018.00</td><td>3.11</td><td>10.05</td><td>16</td><td>0.924</td><td>0.880</td></tr><tr><td><strong>Ericaceae</strong></td><td><strong>1986.95</strong></td><td><strong>1986.95</strong></td><td><strong>7.91</strong></td><td><strong>28.41</strong></td><td><strong>39</strong></td><td><strong>1.000</strong></td><td><strong>1.000</strong></td></tr><tr><td><strong>Oxycoccus sp.</strong></td><td><strong>2013.65</strong></td><td><strong>2017.75</strong></td><td><strong>19.15</strong></td><td><strong>81.28</strong></td><td><strong>45</strong></td><td><strong>1.000</strong></td><td><strong>1.000</strong></td></tr><tr><td><strong>Sphagnum palustre</strong></td><td><strong>2014.75</strong></td><td><strong>2020.35</strong></td><td><strong>10.14</strong></td><td><strong>49.94</strong></td><td><strong>42</strong></td><td><strong>1.000</strong></td><td><strong>1.000</strong></td></tr><tr><td>Drosera rotundifolia</td><td>2020.05</td><td>2020.35</td><td>13.72</td><td>56.14</td><td>6</td><td>0.984</td><td>0.928</td></tr></tbody></table>

Note: Bold text highlights taxa that are both pure and reliable.

Abbreviations: Ienv.cp, time change point based on the indicator value score; IndVal, indicator value score for an individual species; Purity, proportion of bootstrap replicates assigning a taxon to the same category; Reliability, proportion of bootstrap replicates with a *p*-value < 0.05 for a given taxon, indicating a predictable response to the time gradient; Z score, Z score value for an individual species; Zenv.cp, time change point based on the Z score.

around 1990 (to ~4cm), which persisted in subsequent years, S. medium/divinum (z-) gradually replaced it. Although both species were allocated to the z- group, their estimated DWT change points differ strongly (S. cuspidatum: -2.12cm, S. medium/divinum: 17.50cm), indicating a shift within the z- group toward a species less dependent on high moisture levels. A similar transition within the same z group was observed at Fochteloer Veen. Here, DWT fluctuations over centuries led to the dominance of S. russowii (z+), which was interrupted by a two-century-long period of S. medium/divinum (z-) dominance, during which the estimated DWT decreased by a few centimeters. After 1950, the

DWT value increased and stabilised at a level of around 25cm below the surface, which was followed by the dominance of S. palustre (z+) and a noticeable increase in vascular plant presence, including Oxycoccus sp. (z+) and Drosera rotundifolia (z+). Although both S. russowii and S. palustre belong to the z+ group, the transition from a species with a lower to a higher DWT change point (2.25-22.42cm, respectively) suggests a long-term deterioration of hydrological conditions in recent decades.

Pichlmeier Moor has undergone substantial hydrological changes over the past centuries. Initially characterised by a
<figure><img src="image_6.png" /><figcaption>**FIGURE 4** | Plot showing taxon-specific change points along the time gradient for taxa predominantly recorded in earlier (z−) versus more recent (z+) periods.</figcaption></figure>

relatively low water table (DWT ~24 cm) and dominance of *S. medium/divinum* (z−), the site experienced significant drying (DWT ~36 cm), leading to the significant development of vascular taxa such as *Trichophorum/Eriophorum*. However, in the last few decades, a relative improvement in hydrological conditions (DWT ~28 cm) has led to the re-establishment of *Sphagnum* dominance, specifically *S. fuscum/rubellum* (z+).

At Pürgschachen Moor, the early centuries were characterised by medium and fluctuating DWT values (c. 10 cm below surface) and the dominance of *S. medium/divinum* (z−, estimated DWT change point: 15.53 cm). Around the 17th century, the site experienced a temporary increase in the water table (DWT~−5 cm), coinciding with a shift in dominant *Sphagnum* species to *S. cuspidatum* (z−, estimated DWT change point: −2.12 cm). Over the last two centuries, a continuous decline in water tables (DWT reaching ~15 cm) has been observed, marked by the resurgence of *S. medium/divinum* (z−) and the increasing dominance of *S. fuscum/rubellum* (z+).

# 4 | Discussion

## 4.1 | Critical Thresholds of the Water Table in Ombrotrophic Peatlands

An appropriate water table, measured as DWT, remains a key parameter for the conservation and maintenance of peatlands in a favourable ecological state (Haapalehto et al. 2011; Waddington et al. 2015; Kreyling et al. 2021; Baird and Low 2022; Zak and McInnes 2022). Our research identified two change points for both taxa that benefit from lower and higher DWT levels. These findings indicate a transition zone between the dominance of species classified as moisture-dependent and those more drought-adapted, occurring between 7.2 and 22.1 cm DWT. According to our results, an increase in DWT within this range may promote a shift in plant species composition toward more drought-adapted taxa, with the likelihood of such a transition increasing as the water table drops even further.
<figure><img src="image_7.png" /><figcaption>**FIGURE 5** | Plot showing TITAN sum(z−) and sum(z+) values for all potential time change points. Peaks in sum(z) indicate the point beyond which synchronous declines in the summarised z-score values occur within the given taxon group. Dot size is proportional to the Z score of a given taxon.</figcaption></figure>

The defined critical transition zone is broader yet consistent with previous estimations by Lamentowicz et al. (2019). They identified a DWT range of 8–17 cm, with a tipping point at 11.69 cm, likewise derived from TITAN analysis. Similarly, the guidelines provided by Couwenberg (2011) based on recent GHG emission estimates suggest an optimal DWT of approximately ±10 cm, aligning with values reported elsewhere to support the restoration of previously degraded peatlands (Haapalehto et al. 2011). In contrast, Jassey et al. (2018), based on short-term experimental manipulations (2 years) and subsequent shifts in both plant and fungal communities, estimated the transition zone to lie between 19 and 23 cm DWT—a range close to our estimated tipping point for z+ species. Additionally, long-term analyses of peat accumulation rates across several European peatlands have shown that the optimal DWT level for the highest peat accumulation lies between 5 and 10 cm (Swindles et al. 2025). In summary, based on both our findings and previous research, we propose that the critical transition zone for typical bog plant communities ranges from 15 to 22 cm, above which a rapid species shift may occur. Meanwhile, a DWT of 10 cm could be recommended as optimal and safe from the perspective of peatland restoration and conservation.

Empirical evidence supporting these values was also found in macrofossil analyses. For instance, records from Drebbersches Moor and Amtsvenn-Hündfelder Moor documented a rapid turnover in plant species composition in response to increasing DWT. Both sites were initially dominated by *Sphagnum* species from the z− group (such as *S. cuspidatum*, *S. austinii* or *S. medium/divinum*) until the DWT exceeded 20 cm, leading to a significant shift in dominant moss species to those from the z+ group (*S. fuscum/rubellum*) as well as the introduction of a

higher proportion of vascular plant species (*Erica tetralix* and shrubs). Another case supporting the estimated values, where an increase in water table (lower WTD) promoted a shift in the opposite direction, was observed in the partially regenerated Bagno Kusowo peatland. At this site, hydrological improvements led to a DWT change from ~20 to 12 cm. Following this shift, a turnover in the dominant *Sphagnum* taxa occurred, with the more drought-adapted *S. fuscum* (z+ taxon) being replaced by *S. medium/divinum* (z− taxon). While these observations are based on individual case studies, they support the estimations from our research and previous studies.

It is essential to note the similarity between the results of our analysis and those of Lamentowicz et al. (2019), who, for their DWT tipping point estimation, also used the TITAN analysis on paleoecological data obtained from peat core samples. However, in their work, the data—although originating from multiple and diverse peatlands—came from one European country, raising the question of whether these results could be applied to peatlands at the continental or even global level. The fact that our study, which utilised the same TITAN method, but with data from seven distinct sites across the European continent (with only one site shared with Lamentowicz et al. 2019, for which we present original and independent data), yielded similar results further strengthens the reliability of the estimated community thresholds. In addition, it is worth noting that the peatlands examined in our study differ somewhat from those of Lamentowicz et al. (2019), as all of our sites have experienced significant historical drainage and peat excavation and are currently undergoing restoration processes.

It should be emphasised that a species shift caused by a specific decline in the water table does not necessarily imply a loss of
<figure><img src="image_8.png" /><figcaption>**FIGURE 6** | Legend on next page.</figcaption></figure>
**FIGURE 6** | Diagrams showing the proportion of dominant macrofossil taxa in the analysed peat core layers, along with estimated water table changes and shifts in the ratios of z+ (wet-associated taxa—green) and z− species (dry-associated taxa—brown). Negative values of the DWT parameter indicate a water table below the surface, while positive values represent water accumulation above ground level. The full set of the age-depth models is presented in Figures S1–S7, complete plant macrofossils diagrams are presented in Figures S8–S14, and complete testate amoebae diagrams are presented in Figures S15–S21.

peatland functional composition or ecological value. A comprehensive analysis of 56 European peatland sites by Robroek et al. (2017) examined species composition across different environmental gradients and found that while climatic factors may drive species turnover, the overall functional composition of these ecosystems often remains largely unchanged. This can be explained by the fact that disappearing species are replaced by taxa that, although better adapted to new environmental conditions, still fulfil similar ecological roles. These findings suggest that peatlands exhibit a degree of resilience to environmental changes, including those driven by climate change. Nevertheless, a larger increase in DWT will unavoidably translate into an overall net emission of carbon to the atmosphere and thus destroy the peatland C sink function (Evans et al. 2021). Irrespective of this, peatlands that have undergone significant species turnover in the past should still be considered important ecosystems, with their continued conservation remaining a priority despite the considerable transformation.

However, our study provided a contrasting view of the complex histories of peatlands, and a clear look at the real state of the disturbance is only possible with the palaeoecological approaches. Considering all studied sites, we can state that all of them are highly disturbed; however, despite that, the data provide a strong argument for the increased effort of the restoration. New species coming after drainage and exploitation are not comparable to the pristine baseline, as already described for fen systems (Kreyling et al. 2021). Therefore, the peatland baseline-disturbance state relation cannot be assessed only by looking at the peatland surface.

## 4.2 | Time

Currently, a significant portion of the world's peatlands is classified as degraded. In countries historically rich in peatlands, such as Germany, the Netherlands, Poland or Finland, a substantial proportion, partially > 90%, of these ecosystems has been lost or severely altered (Tanneberger, Abel, et al. 2021). Our TITAN analyses, combined with peat macrofossil and DWT reconstructions, provide further valuable insights into the history of peatland degradation across Europe. The findings suggest a pronounced shift in taxonomic composition over time, particularly during the 19th and 20th centuries, with a marked acceleration after 1950—likely in response to increasing ecosystem degradation. This interpretation aligns with historical records (Lamentowicz et al. 2015), which indicate that extensive peat extraction took place from the late 18th century through the first half of the 20th century, typically following drainage efforts. Furthermore, testate amoeba data from multiple European sites suggest that the period between 1800 and 2000 was the driest in the long-term history of these ecosystems (Swindles et al. 2019), reinforcing the hypothesis that significant shifts in

species composition occurred during this time. Following World War II (1939–1945 CE), the exploitation, drainage, and degradation of peatlands intensified across several European countries, leading to substantial increases in DWT (Swindles et al. 2019). In some cases, these processes were so intensive that, for example, in Finland, estimates suggest that up to 70% of the country's mires were lost due to ambitious drainage projects, which reached their peak in the 1970s (Tanneberger, Abel, et al. 2021). This large-scale increase in DWT may be reflected in our second TITAN analysis, which shows along with an increase in DWT a sharp decline in sum(z−) species and an exceptionally rapid increase in sum(z+) species after 1950. Furthermore, the analysis identified a notable shift towards vascular plants, particularly low water table-related taxa such as *Oxycoccus* sp., Ericaceae, *Pinus sylvestris*, Poaceae, and *Drosera rotundifolia*. Similarly, *Sphagnum* taxa (*S. palustre*, *S. papillosum*, and *S. fuscum/rubel-lum*), while still present in the z+ group, included two species that were classified in the first TITAN analysis as part of the z+ group (associated with increases in DWT), indicating a shift toward more drought-tolerant *Sphagnum* species.

Overall, our findings suggest that not only a significant turnover in species composition occurred over the past 200 years, with an acceleration in recent decades, but also that this shift reflects a long-term decline in water availability (Giese et al. 2025). The dominance of drought-adapted species in degraded peatlands supports the hypothesis that hydrological conditions have progressively deteriorated, particularly in the latter half of the 20th century. This aligns well with global wetland change reconstructions (Fluet-Chouinard et al. 2023), which indicates that since 1700, nearly 50% of northern European wetlands, including peatlands, have been lost, with a significant portion disappearing in the second half of the 20th century.

## 4.3 | Implications for Restoration

In recent years, increasing attention has been paid to the risks associated with leaving degraded peatlands unmanaged (Günther et al. 2020; Tanneberger et al. 2020; Tanneberger, Moen, et al. 2021; Zak and McInnes 2022). Globally, it is estimated that even 15% of the total peatland area has been drained (Joosten and Clarke 2002) with such sites potentially accounting for as much as 5% of global anthropogenic CO₂ emissions (Joosten and Clarke 2002; Friedlingstein et al. 2019), not to mention the emission of other important GHGs such as CH₄ or N₂O. Emissions from these degraded ecosystems represent a substantial challenge in the fight against climate change, particularly in Europe, which is the biggest GHG emitter from degraded peatlands just after Indonesia (Joosten 2010). Historically, the continent was extensively covered by peatlands, but it is now estimated that approximately 25% of Europe's peatlands remain degraded, with this figure being as high as 50% within the European Union
(Tanneberger, Abel, et al. 2021). The costs of leaving peatlands in an unsatisfactory state are therefore considerable, making their restoration a crucial task for reducing GHG emissions.

Fortunately, research and practical experience demonstrate that peatland restoration is feasible (Zak and McInnes 2022). While the initial stages of rewetting of degraded peatlands are associated with substantial methane emissions, rewetting has been shown to rapidly and significantly reduce CO₂ emissions (Huth et al. 2022). Although methane's impact on global warming is considerable, as Günther et al. (2020) highlighted in their work, its atmospheric lifespan is much shorter than that of CO₂ (approximately 12 years compared to 100 years). Therefore, in the long term, rewetting a degraded peatland is always a more favourable solution than leaving it unmanaged, as it increases the amount of C accumulated in the peat (Günther et al. 2020; Tanneberger, Moen, et al. 2021). Moreover, studies suggest that methane emissions during the rewetting process can be mitigated. For example, the removal of the degraded top layer of peat during topsoil removal has been shown to significantly reduce methane emissions (Zak and McInnes 2022). In addition, it should be noted that both conservation and restoration of peatlands support most of the United Nations Sustainable Development Goals, such as 'no poverty', 'clean water and sanitation', 'climate action', or 'life on land' (Tanneberger et al. 2020). Thus, while restoring proper hydrological conditions is clearly a priority, the question remains as to what level of DWT is optimal for facilitating the regeneration of degraded peatlands and ensuring the long-term preservation of still-intact peatlands.

Based on our palaeoecological findings and available data, we support the notion that a safe DWT for peatland regeneration is approximately 10 cm below the surface. Higher DWTs are associated with an increased likelihood of shifts in species composition towards drought-adapted taxa. Accordingly, our results suggest that the DWT increasing to the range of ~15–22 cm increases the risk of significant species turnover, thus marking this as a transition zone towards more drought-resistant species. While these values offer essential insights into the relationship between water table and plant communities, their application requires caution. Different peatland sites may exhibit varying tolerance to drainage, resulting in site-specific critical DWT thresholds for bog plant communities. Nevertheless, maintaining the DWT at approximately 10 cm below the surface appears to be a safe target, as it is slightly shallower than the estimated turnover range. Its simplicity as a round number may also facilitate its use in restoration practice and contribute to sustaining healthy peatland ecosystems.

What is worth noting is that the DWT level of around 10 cm is also an essential threshold from the perspective of mitigating carbon emissions from peatlands. Such conclusions regarding carbon emissions were made by Evans et al. (2021), who analysed the relationship between peatland DWT levels and carbon fluxes using the data collected from multiple eddy covariance flux towers located near English and Irish peatlands. Their study revealed that higher DWT levels were associated with substantial CO₂ emissions, whereas lower DWT levels promoted increased CH₄ emissions. However, what is worth noticing is that a DWT level of approximately 10 cm facilitated net CO₂ uptake from the atmosphere, effectively offsetting the CH₄ emissions

observed at this DWT level. This balance should result in an overall absorption of GHG, highlighting the potential importance of maintaining such low DWT levels to mitigate climate change. These findings are consistent with the observations of Fortuniak et al. (2021), who, based on 6 years of continuous eddy covariance CO₂ and CH₄ flux measurements from a single site, concluded that to mitigate GHG emissions, the DWT should not fall below 15 cm. Similarly, Chimner et al. (2017), in their analysis of the long-term effects of water table manipulation on GHG emissions, reported that a site with a DWT of approximately 12 cm acted as the highest carbon sink compared to areas with a lower water table (higher DWT). Thus, both these studies and our own results suggest that a DWT of around 10 cm may be the most optimal for both peatland preservation and carbon emissions mitigation. Such a DWT is likely to not only preserve the ecosystem's integrity or stimulate the regeneration of degraded peatlands but also, within a relatively short period, contribute to efforts to slow the growth of GHG concentrations in the atmosphere. This is particularly important, as many drained peatlands currently act as significant carbon emitters (Leifeld et al. 2019; Doelman et al. 2023).

Aiming for a DWT of 10 cm as a threshold may seem at first sight a challenging task. Moreover, we see water table levels in peatlands varying both over space and time. However, in this regard it is essential to recall that intact peatlands have a set of self-stabilising mechanisms to maintain a stable hydrology (Couwenberg et al. 2022). Although their magnitude and significance are under debate, it is evident that particularly under disturbance, several hydrological feedbacks may become active or more pronounced, such as, e.g., increased evapotranspiration due to ingrowth of trees, or altered hydrological properties of degraded peat, and may further complicate the re-establishment of optimal DWT (Kopp et al. 2013; Waddington et al. 2015). Therefore, while during the restoration of peatlands with hydrologically non-optimal peat properties and/or vegetation initially more effort may be needed to keep optimal DWT, pristine or near-natural peatlands after restoration have the capability to stabilise and buffer water table levels more effectively. The level of necessary human intervention may thus be expected to decrease in restored sites over time, as the hydrological properties increasingly approach pre-disturbance conditions (Ketcheson and Price 2011).

It is crucial to emphasise that, while peatland restoration is achievable, it is neither a straightforward process nor a guaranteed success, as it heavily depends on a thorough understanding of the local hydrological regime (Zak and McInnes 2022). Therefore, priority should be given to protecting peatlands that remain in good condition before they degrade. Our analysis of peat cores indicates that even a relatively short time of water regime disruption can result in significant shifts in peatland species composition. This rapid response to hydrological changes has also been documented in other studies (Couwenberg 2011). In contrast, practical experience demonstrates that peatland restoration is a much more prolonged and complex endeavour, often taking decades to achieve meaningful results (Kreyling et al. 2021). Restoration efforts are typically resource-intensive and costly, while the degradation of a peatland can occur within just a few years of unfavourable conditions. Furthermore, even successfully restored peatlands often exhibit significant differences in biodiversity and species composition compared to those
that have never been degraded, as has been described for fen
systems (Kreyling et al. 2021). Thus, while the functionality
of peatlands may be restored to a certain degree, their pristine
character may never be recovered in a matter of a few decades
(Haapalehto et al. 2011). This issue is also evident in the case
of testate amoebae data from Bagno Kusowo. Although the site
has undergone a self-rewetting process in recent decades, its
testate amoeba community remains significantly different from
the typical species composition observed in similar but non-
degraded parts of the same site. Consequently, while restoration
is achievable, it is a challenging and time-consuming process,
highlighting the importance of proactive protection measures to
prevent degradation in the first place.

Unfortunately, only a limited proportion of peatlands are cur-
rently subject to any form of formal protection, and even among
those that are protected, many faces inadequate safeguards
(Tanneberger, Abel, et al. 2021). For instance, in some cases,
peatlands are included within the boundaries of protected areas,
but they are not the primary focus of conservation efforts. In
other instances, the protection is confined to the peatland body
itself, leaving the catchment in which the peatland is embedded
unprotected, which may threaten the hydrological stability of the
area, essential for peatland preservation; for example lateral and
vertical water losses may occur. Therefore, the authors of this
study call for addressing these gaps in conservation strategy and
applying proper rewetting strategies as a vital step to ensuring
the long-term integrity and functionality of these ecosystems.

5 | Conclusions

Our integrated analysis of testate amoeba-derived DWT recon-
structions and plant macrofossil records identified a critical
transition zone between 7 and 22cm below the peat surface.
Crossing this threshold is associated with significant shifts in
plant species composition, including a decline in *Sphagnum*
taxa, which likely signals the onset of peatland degradation.
Testate amoeba data from Bagno Kusowo further reveal that,
even following hydrological restoration, microbial assemblages
differ markedly from those found in undisturbed, pristine peat-
lands. This suggests that biological recovery may lag behind
hydrological improvements or follow a different trajectory
altogether, although known internal hydrological feedbacks
also complicate restoration and maintaining optimal DWT.
Moreover, analysis of vegetation dynamics across the studied
sites indicates a pronounced shift from moisture-dependent to
drought-adapted plant species beginning in the early 18th cen-
tury and accelerating in the latter half of the 20th century—cor-
responding with increased anthropogenic pressure and a widely
documented increase in DWT. Based on our multi-proxy data
and long-term ecological trajectories, as well as evidence from
previous studies, we recommend maintaining a mean DWT of
approximately 10cm below the surface as a safe and optimal tar-
get for both peatland protection and restoration efforts.

Author Contributions

**Mateusz Draga:** data curation, formal analysis, investigation, visualization, writing – original draft. **Mariusz Gałka:** formal analysis,

investigation, visualization, writing – review and editing. **Klaus-Holger Knorr:** investigation, writing – review and editing. **Stephan Glatzel:** investigation, writing – review and editing. **Bogdan H. Chojnicki:** writing – review and editing. **Christian Fritz:** investigation, writing – review and editing. **Vincent E. J. Jassey:** writing – review and editing. **Hanna Meyer:** writing – review and editing. **Bjorn J. M. Robroek:** writing – review and editing. **Carrie L. Thomas:** writing – review and editing. **Mariusz Lamentowicz:** conceptualization, formal analysis, investigation, supervision, visualization, writing – review and editing.

Acknowledgements

This research was funded through the 2020-2021 Biodiversa+ and Water JPI joint call for research projects, under the BiodivRestore ERA-NET Cofund (GA N° 101003777), with the EU and the funding organisations National Science Centre (Poland) proposal number—2021/03/Y/ ST10/00093 (Poland), DFG (Germany), FWF (Austria), and the Ministry of LNV (The Netherlands). The project 'ReVersal' was coordinated by Klaus-Holger Knorr (Germany) and Stephan Glatzel (Austria) with partners, Bjorn Robroek (Netherlands) and Mariusz Lamentowicz (Poland). The authors are grateful to Patryk Fiutek for the help in the field and for analysing testate amoebae.

Funding

This work was supported by Narodowe Centrum Nauki (101003777).

Conflicts of Interest

The authors declare no conflicts of interest.

Data Availability Statement

The data supporting the findings of this study are openly available in Mendeley at https://doi.org/10.17632/rfnzk5yn8t.1.

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**Supporting Information**

Additional supporting information can be found online in the Supporting Information section. Data S1: gcb70629-sup-0001-FigureS1-S21-TableS1.pdf.