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# Experimental assessment of tundra fire impact on element export and storage in permafrost peatlands

Daria Kuzmina, Artem Lim, Sergey Loiko, Oleg Pokrovsky

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Daria Kuzmina, Artem Lim, Sergey Loiko, Oleg Pokrovsky. Experimental assessment of tundra fire impact on element export and storage in permafrost peatlands. Science of the Total Environment, 2022, 853, pp.158701. 10.1016/j.scitotenv.2022.158701. hal-04908294

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1 Experimental assessment of tundra fire impact

2 on element export and storage in permafrost peatlands

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4 Daria Kuzmina¹, Artem G. Lim¹, Sergey V. Loiko¹, Oleg S. Pokrovsky²,³*

5 ¹ BIO-GEO-CLIM Laboratory, Tomsk State University, 36 av. Lenina, Tomsk 634004, Russia

6 ² GET UMR 5563 CNRS University of Toulouse (France), 14 Avenue Edouard Belin, 31400

7 Toulouse, France

8 ³ Federal Center for Integrated Arctic Research, Institute of Ecological Problem of the North, 23

9 Nab Severnoi Dviny, Arkhangelsk 163000, Russia

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13 Submitted to STOTEN, 29 June 2022. After revision, September 7, 2022.

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17 Key words: moss, lichen, peat, ash, nutrient, river

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19 *corresponding author email: oleg.pokrovsky@get.omp.eu

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Abstract

Extensive studies have been performed on wildfire impact on terrestrial and aquatic ecosystems in the taiga biome, however consequences of wildfires in the tundra biome remain poorly understood. In such a biome, permafrost peatlands occupy a sizable territory in the Northern Hemisphere and present an extensive and highly vulnerable storage of organic carbon. Here we used an experimental approach to model the impact of ash produced from burning of main tundra organic constituents (i.e., moss, lichen and peat) on surrounding aquatic ecosystems. We studied the chemical composition of aqueous leachates produced during short-term (1 week) interaction of ash with distilled water and organic-rich lake water at 5 g solid L⁻¹ and 20 °C. The addition of ash enriched the fluid phase in major cations (i.e., Na, Ca, Mg), macro- (i.e., P, K, Si) and micronutrients (i.e., Mn, Fe, Co, Ni, Zn, Mo). This enrichment occurred over less than 2 days of experiment. Among 3 studied substrates, moss ash released the largest amount of macro- and micro-components into the aqueous solution. To place the obtained results in the environmental context of a peatbog watershed, we assume a fire return interval of 56 years and that the entire 0-10 cm of upper peat is subjected to fire impact. These mass balance calculations demonstrated that maximal possible delivery of elements from ash after soil burning to the hydrological network is negligibly small (< 1-2 %) compared to the annual riverine export flux and element storage in thermokarst lakes. As such, even a 5-10 fold increase in tundra wildfire frequency may not sizably modify nutrient and metal fluxes and pools in the surrounding aquatic ecosystems. This result requires revisiting the current paradigm on the importance of wildfire impact on permafrost peatlands and calls a need for experimental work on other ecosystem compartments (litter, shrubs, frozen peat) which are subjected to fire events.
## Introduction

One of the biggest unknowns about climate warming in high latitudes is the response of wildfire area, fire frequency and fire severity (Flannigan et al., 2000; Kasischke et al., 2005, 2010; Ludwig et al., 2018). The global-scale effect of wildfires in the boreal and subarctic regions stems from large stocks of combustible materials on the forest floor, (i.e., plant litter, mosses and lichens) (Czimzcik et al., 2003; Neff et al., 2004; Boby et al., 2010). For example, fires are the main destabilizing factor in the boreal belt of Siberia (Kasischke and Turetsky, 2006; Kharuk et al., 2021; Talucci et al., 2022). It has been documented that tundra fire is among the most important driving forces of permafrost carbon release into the atmosphere (Miner et al., 2022); this is especially true for the Western Siberian Lowland (WSL), a territory of forest, forest-tundra and peatlands that contains a vast amount of organic carbon (OC) in the form of peat and vegetation (Botch et al., 1995; Beilman et al., 2009).

Through consuming a sizable amount of ecosystem OC stock, wildfires accelerate the recycling of major and micronutrients such as K, S, N, P, Si and trace metals (Whelan, 1995; Moilanen et al., 2002; Spencer et al., 2003). Pyrogenic material (i.e., plant and topsoil ash) produced from the burning of organic matter constitutes a sizable pool of highly reactive materials which can profoundly affect soil physical and chemical properties. The magnitude of these impacts depends on temperature and duration of combustion, soil moisture content at time of burning and identity and chemical composition of the burning vegetation (Wright and Bailey, 1982; Grey and Dighton, 2006; White-Monsant et al., 2015). Tundra fires are known to strongly impact the carbon and element cycling in high latitudes (de Groot et al., 2013; Hanes et al., 2019; Chen et al., 2021). However, for application in forestry and available tree stand assessment, the overwhelming majority of experimental and natural studies in Siberia have dealt with forest rather than tundra vegetation (Ivanova et al., 2011; Larjavaara et al., 2017; Köster et al., 2018; Prokushkin et al., 2018; Knorre et al., 2019). Numerous studies on environmental impact of the tundra fire in other regions (Hu et al., 2015; Burd et al., 2018; Gibson et al., 2018; Heim et al.,
2021) do not assess specific chemical mechanisms responsible for fire product leaching to surface waters. Moreover, relatively little is known about this important consequence of tundra wildfire with potential to modify the chemical composition of surrounding surface waters (thaw ponds and thermokarst lakes) and riverine element export from the watershed (Parham et al., 2013).

Ash produced during vegetation burning is rich in essential nutrients like nitrogen, sulfur, phosphorus and potassium (Koyama et al., 2010; Harden et al., 2004; Neff et al. 2004). Because most of these elements are present in a water-soluble state (Pereira et al., 2012), they are highly mobile over the first days to weeks after a wildfire has occurred (Christensen 1973; Grier 1975; Kauffman et al., 1993). Fire-induced nutrient losses from terrestrial ecosystems with stream runoff (Bayley et al., 1992; Lamontagne et al., 2000; Petrone et al., 2007) are likely to be particularly important in permafrost-dominated regions with shallow soils and short residence time of water at the watershed (Ala-aho et al., 2018). In such regions, wildfires are capable of altering both stream chemistry and water column biological activity (Kawahigashi et al., 2011; Parham et al., 2013, Diemer et al., 2015, Rodriguez-Cardona et al., 2020).

This study aims to provide quantitative constraints on tundra vegetation (i.e., mosses and lichens) and peat ash reactivity in model aqueous solutions; these solutions imitate natural rainfall conditions and the organic-rich nature of surface water in permafrost peatlands. As a working hypothesis, we suggested that moss ash exhibits the highest reactivity in aqueous solutions. This hypothesis is based on recent results of mesocosm experiments with live vegetation in permafrost peatlands (Shirokova et al., 2021; Manasypov et al., 2017). A secondary hypothesis, in accord with the existing paradigm that the wildfire is among the most significant factors for climate warming impact on the biogeochemical cycle of elements in the permafrost regions (Abbott et al., 2016, 2021; Ackley et al., 2021), was that tundra wildfire is capable of strongly modifying element transfer between burned peat/soil and adjacent aquatic systems. To assess this, we compared the lateral export (yield) of major and trace elements from burned
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mosses, lichens and peat at the watershed scale to the annual riverine yield of nutrients, metals and dissolved element storage in thermokarst lakes.

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Quantifying the magnitude of tundra vegetation and peat ash reactivity in naturally-relevant solutions should allow for a better prediction of the scale of impact that tundra wildfires have on the surrounding hydrological network. This information is greatly needed to foresee changes in element delivery from land to ocean amidst on-going climate changes and environmental disturbances in western Siberia and other global permafrost peatlands.

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## 2. Methods

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### 2.1. *Moss, lichen and peat ash preparation*

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Sampling of peat, vegetation and lake water was performed in August 2020 at the Khanymey Research station (INTERACT Network), 63°47′N, 75°43′E. The upper peat horizon used in this study represented the first 0-10 cm of soil after removal of moss and litter. This is the typical peat depth affected by tundra fire according to Liljedahl et al. (2007). It was sampled at the peat mound and dried at 40°C using Osmofilm® bags over 5 days. The peat was 5 to 10% decomposed and contained 55% lichens (40% of *Cladonia stellaris* and 15% of *Cladonia stygia*), 30% dwarf shrubs, 15% Sphagnum moss with some wood remnants, bark and pollens as determined by botanical analysis (Morgalev et al., 2017). The peat was slightly ground/disintegrated before subjecting to ashing. The apical (live) part of *Cladonia stellaris* and *Sphagnum fuscum*, highly abundant in the frozen mounds of the palsa peat bog at the Khanymey Research Station and constituting > 90% of peat mound surface vegetation (Volkova et al., 2018), were rinsed in Milli-Q water and dried in open air using Osmofilm® bags.

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Approximately 80 g of dry peat, lichen or moss were placed in pre-cleaned porcelain cups and separately burned for 3 hours at 525 °C in a Vecstar® furnace. This temperature corresponds to the upper range of in-situ temperatures recorded during tundra ground fire (Agafontsev and Kasymov, 2020) and was chosen to ensure maximal burning of organic material
thereby representative of an extreme case of wildfire. The motivation for this choice was that, unlike organic-rich soils of flat peatlands subjected to smoldering (i.e., Benscoter et al., 2011), the permafrost palsa peatbogs investigated in the present study are represented by convex mounds that do not hold water and become extremely dry during summer time. The ash produced by this burning had a black color potentially indicating not a 100% removal of OM (i.e., Bodi et al., 2014), as was also reflected in the small amount of residual organic carbon (mass loss 1 to 2.5 % after burning at 1200 °C). Note that recently, Li et al (2022) demonstrated incomplete consumption of organic matter even at 525 °C. The ash mass content of dry peat, lichen and moss biomass was equal to 2.5±0.5, 0.8±0.1 and 1.8±0.7 %, respectively, as measured by triplicate analysis.

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## 2.2. Laboratory experiments on ash dissolution

For leaching experiments, sterile 0.22-µm filtered and UV-treated Milli-Q water and natural lake water were used. The lake water was collected in August 2020 from the surface of a thermokarst lake located at the Khanymey Research Station (11,000 m² surface area, depth 0.5-1.0 m) and immediately filtered through a sterile 0.45 µm Nalgene filtration unit and then stored at 4°C prior the experimentation.

Ash leaching experiments were conducted at a fixed concentration of solid phase (5.0 g L⁻¹) in sterile, pre-cleaned polypropylene bottles, in triplicates. The bottles were gently stirred on a ping-pong shaker (100 rpm) at 20±0.5°C in darkness. One series of experiments was conducted with Milli-Q distilled water and another with natural organic-rich lake water (DOC of 40 mg L⁻¹). The choice of these two media was to model the two main scenarios of vegetation ash behavior in natural systems after massive fire events: 1) leaching by rainwater and 2) transfer of ash particles to organic-rich surface waters (notably to soil depressions and small thaw ponds).

After 1, 5, 10, 30, 100, 300, 1440, 3000 and 10020 minutes of reaction, the replicate bottles were removed from the shaker and the contents of each bottle were entirely used for
chemical analyses. Aqueous suspensions were filtered immediately after sampling using a sterile
syringe and a disposable, single-used 0.45 µm Millipore filter unit.

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## 2.3. *Chemical analyses*

The pH and electrical conductivity were measured in unfiltered waters immediately after
sampling using Hanna and WTW portable instruments. Dissolved organic carbon (DOC) and
major and trace elements were measured in 0.45-µm filtered samples after their storage of 5 to 7
days in the refrigerator. The major anion concentrations (Cl⁻ and SO₄²⁻) were analyzed by ion
chromatography (Dionex 2000i) with an uncertainty of 2%. The international standards ION,
PERADE, and RAIN were used for validation of analyses. DOC and dissolved inorganic carbon
(DIC) were determined by a Shimadzu TOC-VSCN Analyzer with an uncertainty of 3% and a
detection limit of 0.1 mg/L. Specific UV absorbance (SUVA) was measured as absorbance at
254 nm normalized for DOC concentration in L mg⁻¹ m⁻¹. Major cations (Ca, Mg, Na, K), Si and
~40 trace elements were determined with an Agilent iCap Triple Quad (TQ) ICP MS using both
Ar and He modes to diminish interferences. About 3 µg L⁻¹ of In and Re were added as internal
standards along with 3 various external standards. Detection limits of TE were determined as 3x
the blank concentration. The typical uncertainty for elemental concentration measurements
ranged from 5-10 % at 1-1000 µg/L and 10-20 % at 0.001 – 0.1 µg/L. The Milli-Q field blanks
were collected and processed to monitor for any potential contamination introduced by our
sampling and handling procedures. The SLRS-6 (Riverine Water Reference Material for Trace
Metals certified by the National Research Council of Canada) was used to check the accuracy
and reproducibility of analyses (Heimburger et al., 2013). Only those elements that exhibited
good agreement between replicated measurements of SLRS-6 and the certified values (relative
difference < 15%) are reported in this study.

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Chemical analysis of ash constituents was performed using an X-ray Fluorescence
spectrometer (X Bruker S2), via fusing with Li borate at 1200 °C (XRfuse6, Bruker) and using a
set of sediment and rock standards to quantify major elements (i.e., Na, Mg, Al, Si, P, S, K, Ca, Mn, Ti, Fe) with an average uncertainty of 1 %. Concentrations of these elements were in agreement with certified values of geostandards and were within 10%. Morphology and chemical composition of ash particles were also examined via SEM (Scanning Electron Microscopy) using a Tescan Vega 4 LMU electron microscope equipped with an EDX (energy-dispersive X-ray) detector of SDD Bruker Quantax 30mm².

## 2.4. Statistical tests

Experimental parameters which included pH, temperature, and all measured major and trace element concentrations were analyzed with best fit functions. The best fit functions were based on the Pearson correlation and a parametric ANOVA (Analysis of Variance) on samples where the Shapiro-Wilk test yielded the normality of the data and the Barlett test demonstrated homogeneity of variance. Regressions and power functions were used to examine the relationships between pH, DOC, and major and trace element concentrations as a function of leaching time. The criterion for significant correlation between the chemical parameters was $p < 0.05$. If the ANOVA was significant at $p < 0.05$, a parametric Student t-test and Tukey test were used given that data were normally distributed. A lower probability ($p < 0.05$) indicated that the overall differences in concentration – time trend are important and statistically significant compared to internal variations between replicates. The pairwise trace element correlations (Cl, SO₄, Si, P, K, Al, Fe, Ca and DIC as major leachable constituents of ash material) in the filtrates over the course of incubation experiments were tested using Pearson correlation (SigmaPlot version 11.0, Systat Software, Inc) for all aqueous solutions and separately for peat, lichen and moss ash at $p < 0.05$. The weight loss of ash in the course of experiment was calculated based on the sum of inorganic solutes Ca, Mg, Na, K, Si, P, Al, Cl, DIC and SO₄, normalized to oxides, in order to monitor the degree of solid phase dissolution. The apparent bulk rates of element removal from ash (mass-normalized elementary yields) were calculated for the first 50 h of
reaction and for the remaining period of exposure (50-167 h). The reason of choosing 50 h as a threshold time for rate calculation is that the majority (> 80-90 %) of ash constituents were leached during first 2 days of experiment.

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## 3. Results

### 3.1. Chemical and mineralogical nature of ash

Ash content of the studied organic substrates followed the order: lichen (0.75 weight %) < moss (1.75 %) < peat (2.5 %). The SEM examination of lichen ash demonstrated that it contains a sizable amount of sand (quartz) grains, isomeric particles of black carbon, and other particles of irregular shape containing K, P, Ca, Si, Mg, Al and Fe (Fig. S1 A). Results of semi-quantitative SEM-EDX examination of lichen ash particles were consistent with the data of total chemical analysis performed by XRF and demonstrated a high proportion of Si (88 % SiO₂), followed by Al (3.2 % Al₂O₃), K (2.0 % K₂O), Ca (1.7 % CaO) and Fe (1.1 % Fe₂O₃), Table S1. Moss ash demonstrated the presence of clearly identifiable phytoliths and diatom frustules, black spheres of amorphous Si 5 to 10 µm in diameter, and various irregular particles with typical elements (other than C and O) detected by EDX being Si, Al, K, Fe, Ca, Mg, Na, P (Fig. S1 B). Total chemical analysis of moss ash demonstrated, in addition to Si, high concentrations of S (15.3 % SO₃), K (1.5 % K₂O), Ca (14.7 % CaO), Mg (9.7 % MgO), P (6.4 % P₂O₅), Na (3.6 % Na₂O) and Fe (2.5 % Fe₂O₃). Peat ash contained vegetation debris and phytoliths, with high concentrations of Si, Fe, Al, Ca and S (Fig. S1 C). Accordingly, total chemical analysis of peat ash showed dominance of Si (43 % SiO₂), followed by Al (11.4 % Al₂O₃), Fe (11.0 % Fe₂O₃), Ca (10.8 % CaO) and S (7.8 % SO₃). XRD analysis of lichen and peat ash demonstrated only quartz peaks whereas moss ash yielded these same peaks and those showing some presence of calcite and dolomite (Fig. S2).

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## 3.2. Element concentrations and yields during ash dissolution

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A foremost important experimental result is the drastic difference in the aqueous reactivity of vegetation ashes. Among the three substrates, moss ash was by far (a factor of 2 to 10) more reactive because it yielded the largest changes in solute concentration over first two days of reaction time (Fig. 1 and Table 1). Thus, moss ash produced sharper change in fluid pH (10 over the first 100 min of reaction, gradually decreasing to 8 after several days of exposure) regardless of the nature of the liquid substrate (Milli-Q water or lake water), as illustrated in Fig. 1. In contrast, peat and lichen ash only slightly modified the pH of the fluid, which decreased from 6 to 5 over the experiments with both Milli-Q and lake water. Similarly, moss ash released a much larger amount of soluble salts into the fluid compared to the other two substrates; the water Electric Conductivity in moss ash experiments ranged from 400 to 600 µS cm⁻¹ compared to 100 to 200 µS cm⁻¹ for peat and lichen ash. Chloride exhibited steady release in both aqueous solutions with maximal effects observed in moss; sulfate concentration, also maximal (2 to 3 times higher) in moss compared to other substrates, remained rather constant after an initial abrupt increase at the very start of experiment (Fig. 1).

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Another experimental result is the very fast change in aqueous solution composition, for both Milli-Q water and lake water runs, so that the concentrations of these constituents were orders of magnitude higher than those of the starting solutions, and this occurred after the first minute of reaction, as can be seen in Figs 1, 2 and 3. We also noted that the total weight loss (in % of the initial mass of solid) became constant after the first hours of reaction following the order: peat (< 1.5 %) ≤ lichen (1.5-2 %) < moss (3 %). The 80 to 90 % of elemental release over first 10 minutes was observed for major elements (SO₄, Si, P, Mg, K and Ca), Al, Fe, micronutrients (V, Cr, Mn, Cu, Zn, Se, Sr, Rb and Mo), trivalent and tetravalent hydrolysates, trace oxyanions and heavy metals.
The temporal pattern of elemental release from three ash samples was generally similar between DOC-rich lake water and distilled water, which allowed generalizing the element concentration evolution among different solid substrates and aqueous solutions. According to the temporal evolution of concentration during experiments, 2 main groups of major and trace elements were identified (Table S2). The 1st group includes major anions such as Cl, SO₄ and DIC, cations such as Na, Mg, Ca and K, macro and micronutrients (Si, P, Co, Mo, Rb and Se) and other trace elements (Ti, Cr, As, Sb, Zr, Y, Cs, Ba, Tl, Th) that steadily increased in concentration over the course of exposure or kept essentially constant [after an abrupt, almost instantaneous increase over the first minute of reaction]. Examples of labile elements in this group (Si, P, Mn, Co, As, Mo, Sb, Rb) are illustrated in Fig. 2. Note that this 1st group can be subdivided into two main chemical behavioral types: 1) Cl-type elements (Cl, B, Mg, Si, P, Ti, Mn, Co, As, Zr, Mo, Sb and Th) gradually increasing their concentration with time and 2) SO₄-type elements (Na, K, Ca, Cr, Mn, Fe, Se, Rb, Sr, Cs and Ba) exhibiting high concentration after the first minute of sampling with a rather constant pattern afterwards.

The 2nd group is formed by those elements exhibiting a systematic decrease in pH value and a decrease in concentrations of Cu, Zn, Cd, Y, LREE and U after the beginning of leaching experiment. In the DOC-rich lake water, DIC and DOC decreased their concentration as illustrated in Fig. 3 (for DOC, Cu, Zn and Cd) and Fig. S3 for V, Y, La and U. In contrast to DOC (Fig. 3 G), the SUVA of the lake water systematically increased in the presence of all three ashes (Fig. 3 H). Some elements demonstrated non-systematic variation over time when reacting with moss and lichen ash, with a local maximum or both maximum and minimum (Cu, V, heavy REEs and Th).

Both elemental groups identified were confirmed by pairwise (Pearson) correlations which simultaneously tested both the aqueous media and the three ashes, as well as individually, for each ash of each material (Table S3). Considering all the solid substrates together, in the leachates, labile elements (B, Na, V, Se, Mo, Rb, Cs, Sb and U) strongly correlated with K ($R_p \ge$
288 0.90; p < 0.01) whereas immobile elements (Ti, Zr, Y and REEs) correlated with Fe (R<sub>P</sub> ≥ 0.80;
289 p < 0.05). This has also been observed individually for moss and peat ash leachate. In the peat
290 ash leachate, Al strongly correlated with low solubility Ti, Cr, Fe, Zr and REEs (R<sub>P</sub> > 0.91; p <
291 0.01) indicating a concomitant release of these elements from (refractory) silicate minerals of
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294 The yields (mass of element released into aqueous solution over a given amount of time
295 and normalized to initial mass of ash) for the first two days (0-50 h) and subsequent 7 days (50-
296 167 h) of leaching in the Milli-Q and lake water for three ash materials (Table 1) are illustrated
297 for some elements in Fig. 4. Exempting Si and Cl, the main release of all elements, including
298 DOC, occurred over the first two days for all ashes with < 20 % of release occurring between
299 subsequent 50 and 167 h of reaction. Phosphorus from lichen ash and Mg from moss ash
demonstrated 20-30 % release during the second stage of reaction. DIC was also sizably (50\%)
300 released during the second stage from moss ash. There was no sizable difference in element
301 release patterns between distilled water and organic-rich lake water (Fig. 4). Release of the
302 largest number of major and trace elements was dominated by moss followed by lichen and peat
303 ash (Table 1). These included DOC and DIC, labile macro- and micro-nutrients (B, Cl, SO4, Na,
304 Mg, K, Mn, V, Zn, Se, Rb, As, Mo, Sb, Cs, W and U) and some trace metals (Al, Cr, Cu, Co, Ni
305 and Cd). The 2nd group of elements, whose yield was dominated by lichen rather than moss and
306 peat, included two major nutrients (Si, P) and a number of low mobility elements such as Fe, Ti,
307 Y, Zr, REEs, Pb and Th. Lastly, the yield of alkaline-earth elements was highest in the peat ash
308 followed by moss and lichen ash. The order of release from peat : moss : lichen was 17:16:1,
309 50:10:1, and 16:6:1 for Ca, Sr, and Ba, respectively.
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3.3. Tundra fire impact on rivers and lakes
The experimental results on short-term aqueous reactivity of the three organic substrates dominating the palsa peatlands in western Siberia (moss, lichen and upper peat layer) allow first-order quantification of tundra fire impact on element mobilization from burnt surface to the hydrological network. Considering the proportion of ash in peat, lichen and moss (2.5±0.5, 0.8±0.1 and 1.8±0.7 weight %, respectively), the average elemental content of water-soluble (leachable) elements in organic matter ash can be calculated as the average yields in Milli-Q water and lake water (Table S4). The fresh (dry) organic mass – normalized element yield preserved the same order ‘moss > lichen > peat’ for most macro and micro-nutrients. Considering the total moss and lichen biomass coverage of typical permafrost palsa peatbogs is 1:1, the typical lateral stock of moss and lichen in the northern part of western Siberia is 290 and 245 $g_{d.w.} m^{-2}$, respectively (Kosykh et al., 2008). For this region, with typical peat density of 0.09 kg $cm^{-3}$, the upper 0-10 cm peat mass stock is 9,000 $g_{d.w.} m^{-2}$ (Novikov et al., 2009).

The frozen peatbog (palsa) surface area that is subjected to annual burning in the north of western Siberia has not been quantified yet but can be approximated as 40 % of the total surface per 60-year period (of which tundra and palsa together account for ¾ of total land area according to Sizov et al. (2021)). Therefore, the northern palsa peatbogs of WSL can be entirely burned each 56 years, which is in agreement with our unpublished results on radiogenic C dating, as well as with the latest estimates for the forest tundra and northern taiga of Central Siberia (Novenko et al., 2022; Dvornikov et al., 2022). Note that the average fire rotation time in Eastern and Central Siberian taiga and tundra was recently estimated at 126 years (Talucci et al., 2022), whereas more conservative estimation suggests entire palsa burning each 150 years (Sizov et al., 2021; Novenko et al., 2022). In order to quantify possible release of elements from burned vegetation and peat, we will consider a flat watershed divide near the Khanymey Research Station, containing 40% of lakes, 28% of palsa peatbogs, 8% of forest and 24% of wet depressions and fens (Novikov et al., 2009). Assuming a fire return frequency of 56 years, the wildfire in this region can burn annually 1450, 1225 and 45000 kg $km^{-2}$ of sphagnum, lichen and
339 0-10 cm peat, respectively. Note that in this scenario, we ignore burning of dwarf shrubs and grasses in the palsa and organic litter and forest floor in the forested regions.
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341 These aerial stocks of moss, lichen and upper peat layers, combined with water-soluble element concentrations in vegetation and peat ash (Table 1, Table S4) allow for quantification of the water-soluble yield of element from the watersheds of permafrost peatbogs. For this, assume that the ash produced during tundra burning is entirely mobilized into hydrological network (Table S5). In such a scenario, we test the maximal impact of tundra wildfire on element transfer from the soil to the aquatic systems. In this table, we separately present the water-labile stock of element in peat, moss and lichen ash as well as integral yield of elements from burning of moss, lichen and upper 0-10 cm peat layer. We further assume that all palsa surfaces are burned each 56 years. These possible annual ash-originated element yields can be compared with actual annual yields of major and trace elements by small and medium sized rivers draining through palsa peatbogs of western Siberia; these have been assessed via hydrochemical and hydrological analyses of more than 30 rivers across the seasons in this region (Pokrovsky et al., 2020). For this investigation, we calculated a ratio of annual element leaching from burned vegetation to annual hydrochemical yields (Fig. 5). Overall, the impact of palsa fire on element export from the fire-affected watershed by rivers is negligible and does not exceed 0.1-0.2 % of annual riverine yield for most major and trace elements. Only the riverine yield of Mo, Rb and Sb could be affected by tundra fire at about 0.5-1 % and that of K, SO₄ and Cs at 2-2.5%.

359 Another mass balance calculation was performed to assess the possible impact of tundra fire on the hydrochemistry of thermokarst lakes that are highly abundant in permafrost peatlands.
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361 For this, we considered two extreme scenarios of the lake coverage in the territory subjected to wildfire: 40% for the Khanymey Research station test site described above and 4 % which represents the minimal limnicity of a much larger (i.e., 1 million km²) permafrost-affected part of the WSL (Polishchuk et al., 2018). The average lake depth in the permafrost palsa peatlands was
assumed to be 1 m (Manasypov et al., 2020). We further assumed that all ash produced by tundra burning will be delivered to adjacent lakes via surface or suprapermafrost flow, which represents maximal possible impact of a wildfire event. The range of major and trace element concentrations in the continuous and discontinuous permafrost zones of the WSL is based on seasonal analysis of more than 70 lakes across a permafrost gradient (Manasypov et al., 2020). These mass-balance calculations demonstrate that the concentration of overwhelming majority of dissolved elements present in the thermokarst lake water column will be negligibly (< 1-2 %) impacted by ash product delivery even in the case of minimal possible limnicity (4 %) of the territory. One exception is Ca, whose concentrations in thermokarst lakes (range 0.2 – 1.2 mg L⁻¹, Manasypov et al., 2020) can be non-negligibly affected by ash leaching in a scenario of low limnicity (addition of 0.18 mg L⁻¹). Finally, only phosphorus originated from ash dissolution (addition of 2 and 20 µg L⁻¹ for watersheds having 40 and 4% of lake coverage, respectively) can sizably modify the current Pₜot concentration in WSL lakes (1 to 3 µg L⁻¹).

Note that the mass-balance calculations presented above have to be considered with caution because we averaged a unique event of total palsa burning (28% of the territory) to the total duration of fire return frequency (56 years). A single-year effect of moss, lichen and upper peat burning on element export by rivers and lake hydrochemical composition might be drastically (up to 56 times) higher. In such a case, the delivery of ash leaching product can at least double most element fluxes and pools in aquatic systems for that particular year.

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## 4. Discussion

### 4.1. *Nature of ash from moss, peat and lichen*

Numerous works have documented the extremely high reactivity of ash formed on the surface of the soil after a fire and its particular enrichment in essential nutrients like N, S, P and K (Koyama et al., 2010; Harden et al., 2004; Neff et al. 2004; Pereira et al., 2012). In contrast to
the vast amount of literature devoted to physico-chemical properties and reactivity of wood and grass ash (Fang and Jia, 2012), the information on chemistry and mineralogy of peat, moss and lichen ash remains rather limited. In peat ash, the dominant Si compounds are known to be quartz with occasional amounts of K and Na feldspar; alternatively, clay minerals in peat provide high Al and Fe contents (Steenari et al., 1999). Calcium is known to occur as an oxide, carbonate or sulfate and Mg and P are not usually identified in crystalline form (Steenari et al., 1999). The results of the present study generally agree with the bulk of available information on ash chemical composition and allow for understanding the relative difference in the leachability of different elements. Thus, the total chemical composition of different ashes (Table S1) followed an order which is similar to the order of element leachability observed in the experiments such as 'moss' > 'peat' > 'lichen' for Na, Mg, Si, P, Mn and 'peat' > 'moss' ≥ 'lichen' for Al and Fe. The moss ash released a factor of 2 to 10 higher amount of alkalis (Na, K, Rb, Cs) and nutrients (P, Mn) compared to other ashes because it contained the highest concentrations of these elements in the solid phase. In contrast, Si was mostly leachable from lichen ash (2 and 6 times higher than from moss and peat, respectively, Table 1), where it was concentrated in the form of amorphous silicate material and quartz grains. Although peat ash contained two times more total Si than the moss ash, the later exhibited about four times higher leachable Si compared to peat ash (Fig. 4). This is presumably due to the abundant pool of reactive silica in the form of phytoliths and diatom frustules (Fig. S1).

Note that, although the resolution of XRD analysis (1-2% (w/w)) generally does not allow detecting crystalline phases in ash (i.e., Steenari et al., 1999), among all three studied substrates, only the moss ash demonstrated some peaks of calcite and dolomite (Fig. S2). We believe that the presence of these highly soluble carbonate minerals containing various trace components is responsible for the highest yield of Mn, Co, V, Ni, Cu and other divalent metals which could be associated with these minerals in the moss ash. Further physico-chemical studies of ash reactivity with high resolution spatial examination of both initial solid phases and the
reaction products, including laser ablation ICP-MS and Transmission Electron Microscopy are needed to reveal chemical and mineralogical controls on element leachability from ash to aqueous solution.

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## 4.2. Mechanisms of element leaching from ash

In this section, we discuss the impact of aqueous substrate and mineral source in the solid phase on patterns of element leaching from ash. Most elements constituting the ash material produced after wildfires are present in a water-soluble state (Pereira et al., 2012) and exhibit extremely high mobility in soils (Christensen 1973, Grier 1975, Kauffman et al., 1993). Previous field mesocosm experiments in thermokarst lakes with various organic and inorganic substrates demonstrated that ash is extremely reactive in natural waters, and > 90% of its elemental stock is released over the first 10 h (Manasypov et al., 2017). High-resolution laboratory experiments of the present study demonstrate even faster release of ash constituents with the majority (> 80%) of labile elements mobilized from the solid phase over the first 10 minutes. Unfortunately, the resolution of our sampling and filtration did not allow testing the kinetics of dissolution over a shorter time period. Furthermore, given that such fast rates of ash dissolution likely reflect diffusional versus surface reaction control (i.e., Schott et al., 2009), alternative techniques of measuring ash reactivity (column-through experiments, rotating disk technique, etc.) are needed to quantify rate constants.

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Only a few elements demonstrated gradual change in concentration. This could be linked to mineral dissolution/precipitation controlled by a gradual pH change. This was especially pronounced for moss ash, whose pH decreased up to 2 units over the course of experiment. We therefore hypothesize that labile elements which gradually increased in concentration over time (Cl, DIC, B, Mg, Si, P, Mn, Co, Mo, As and Sb) were incorporated into or adsorbed on the
443 surface of soluble (amorphous) salts such as carbonates, phosphates, oxides or silicates of Na, K
444 and Ca. The dissolution rate and solubility of these sparingly soluble salts is known to increase
445 (e.g., Schott et al., 2009) with a decrease in pH (from 10 to 8 for moss ash and from 6 to 5 for
446 peat and lichen ash). Therefore, such a pH decrease in the course of present experiments could
447 progressively release the aforementioned elements to the aqueous solution.

448 Note that an enhanced mobility of oxy-anions at high pH was created by moss ash
449 dissolution as this solid substrate demonstrated the highest yield of Mo, Se, Sb, W and U. This
450 observation can be explained by a desorption of these oxyanions from moss ash mineral
451 particles. Among these elements is dissolved U(VI) most likely present in the form of uranyl-
452 carbonate or hydroxyl complexes. In contrast, the decrease of Al concentration after 100 min of
453 reaction observed in moss ash experiments in lake water and Milli-Q water could be due to the
454 solubility threshold of amorphous Al hydroxide starting to precipitate from aqueous solution
455 below pH 9.

456 Another group of labile elements (i.e., SO₄, Na, K, Ca, Cr, Se, Rb, Sr, Cs and Ba)
457 exhibited high concentrations after first minute of ash-aqueous solution reaction followed by a
458 rather constant pattern afterwards. We believe that fast dissolution of sparingly soluble sulfates
459 of alkaline-earth and alkaline metals can be responsible for this behavior. A sizable correlation of
460 highly mobile B, Na, V, Se, Rb, Mo, Cs, Sb with K (and to a lesser degree SO₄; **Table S3**) in
461 leachates from all three ashes may indicate presence of these elements in the form of K
462 hydroxide and sulfate. It is also possible that ash material, formed after biomass burning, was
463 composed of multi-layer amorphous/semi-amorphous solid phases of varying chemical
464 composition and reactivity so that progressive surface layer dissolution could expose the interior
465 of the particle to leaching. This could produce a stepwise element concentration pattern. For
466 example, progressive removal of an external Ca silicate, sulfate or phosphate layer at the
467 beginning of reaction exposed a more soluble inner layer containing Na and Mg chloride. To
further examine this possibility, high-resolution future transmission electron microscopy coupled
with elemental analysis has to be performed on both initial phases and reaction products.

The last group of elements was represented by those showing a systematic decrease in
concentration over time, detectable after the first minute of reaction: pH value, Cu, Zn, Cd, Y,
REE and U. Some elements demonstrated non-systematic variation in the course of time for
reaction with moss and lichen ash, with a local maxima or both maxima and minima (V, Mn, Cu,
Zn, heavy REE, Th). It is possible that a decrease in pH was due to transformation of initial
alkali and alkaline-earth hydroxides into carbonates linked to progressive CO₂ uptake from the
atmosphere. Such a process has been documented in pine ash (Pereira et al., 2012). However, we
could not provide a straightforward explanation for these element temporal patterns and their
attribution to either group of solutes. In particular, co-mobilization / precipitation of low
solubility "lithogenic" elements (trivalent and tetravalent hydrolysates, Pb) could occur in the
form of Al and Fe hydroxide colloids, stabilized by a polymerized (highly aromatic) organic
matter that is released from ash material or already present in natural lake water. This is
confirmed by positive correlations (R<sub>P</sub> ≥ 0.80, p < 0.05) observed among these elements in
experimental leachates (Table S3) and is consistent with observations of organo-ferric organo-
aluminium colloids which act as carriers of lithogenic elements in surface waters of permafrost
peatlands (Pokrovsky et al., 2016). Note that a decrease of lake water DOC and an increase in
SUVA<sub>254</sub> (Fig. 3 G, H) in the presence of all three substrates suggest a sizable adsorption of
essentially aliphatic (non-aromatic) dissolved organic matter on ash minerals. As a result,
aromatic, highly polymerized DOC remains in the leachate and can stabilize Fe-Al-rich colloids.

To summarize, there are multiple physico-chemical processes shaping the aqueous
solution chemistry in the presence of ash material, which include: mineral dissolution and
precipitation, adsorption/desorption and co-precipitation. These processes are highly sensitive to
both the solution pH and major element concentrations because the latter determine the overall
solution saturation state with relevant solid phases. However, under static conditions of batch
494 (closed) reactors used in the present study, one cannot identify specific mechanisms controlling
495 element release rate as a function of time and chemical composition of aqueous solution. To
496 distinguish the impact of solution chemistry (pH, element concentration) and relative reactivity
497 of different solid phases, mixed-flow or column leaching dynamic experiments are necessary.

498

499

500

501

### 4.3. Comparison of reactivity of various ashes

502 Both aqueous solutions used in laboratory leaching experiments – Milli-Q water and
503 DOC-rich natural lake water – exhibited similar element release patterns with the majority of
504 elements mobilized to the aqueous solution within 100 min. Thus, ash produced by tundra
505 burning is highly reactive when in contact with both atmospheric precipitates (rain, snow water)
506 and DOM-rich surface waters (thaw ponds, depressions, suprapermafrost waters). Regardless of
507 the DOC level, there was no sizable difference in pH values of solutions interacting with moss
508 ash (10 to 8) and lichen/peat ash (6 to 5). Because pH is the main factor controlling dissolution
509 of most oxides and carbonates which constitute the ash material (Schott et al., 2009), we
510 therefore conclude that the complexation of trace metal with DOM of lake water did not affect
511 release of these metals from the ash material. This is consistent with results of laboratory
512 experiments, which demonstrated a lack of natural DOM impact on dissolution rates of Ca, Mg
513 oxides, carbonates, and silicates (Pokrovsky et al., 2021 and references therein).

514 Compared to lichen and peat ash, moss ash exhibited the highest yield of most macro-
515 and micro-nutrients. This was due to a high initial concentration of these elements in the solid
516 phase (see section 4.1) and a high pH of the contact solution (which presumably stemmed from
517 high concentration of alkali and alkaline-earth metal hydroxides in moss ash). Furthermore,
518 desorption of oxyanions (P, Mo, As, Sb) from mineral surfaces also increases with the pH of
519 solution (Stumm and Morgan, 1995). Note that under such alkaline conditions, we observed
520 exceptional mobility of Al, but not other trivalent hydrolysates, due to the dominance of
521 Al(OH)₄⁻ soluble hydroxo-complexes at pH levels over 9. Uranium could also become highly
522 mobile in the form of its hydroxo- and carbonate complexes.

523 At the same time, lichen ash demonstrated 5 to 10 times higher release of P, Fe, light
524 REE and Pb and 2-3 times higher release of Zn and Cd compared to moss ash. While
525 atmospherically deposited heavy metals might be preferentially accumulated in lichen biomass
526 compared to moss biomass (Gonzalez et al., 2016; Fortuna et al., 2021), enrichment in
527 phosphorus and trivalent cations may be due to some phosphate minerals accumulating in lichens
528 via cyanobacterial symbionts. The prevalence of alkaline-earth metals (Ca, Sr and Ba) in peat
529 ash leachate (as compared to moss ash leachate) may reflect enhanced reactivity of carbonate
530 minerals at 5-4 units lower pH of the peat ash experiments. This could be due to rapidly
531 dissolving carbonate minerals that incorporated these cations.

532 It is important to note that the highest release of macro- and micronutrients from moss
533 ash—as compared to the lichen and peat ash—is consistent with recent results of field mesocosm
534 experiments where the same 3 main organic substrates (i.e., moss, lichen and peat) of permafrost
535 peatlands were separately added in similar proportions to experimental tanks of thermokarst lake
536 water (Shirokova et al., 2021). These experiments demonstrated that mass-normalized release of
537 DOC, macro-nutrients (Si, N and P), major cations, micro-nutrients and trace metals were
538 essentially controlled by moss and, to a lesser degree, lichen with peat having almost negligible
539 impact. Shirokova et al. (2021) concluded that in contrast to the dominant view of the primary
540 role of peat thaw in controlling element biogeochemical cycles in permafrost regions in the
541 context of climate warming, the change in vegetation coverage [notably moss/lichen
542 competition] may be a governing factor for element delivery from land to aquatic systems.
543 Results of this work confirm the overwhelming importance, on a mass-normalized scale, of moss
544 ash reactivity in providing macro- and micro-nutrients to surface waters.

545
546

4.4. Weak impact of peatland fire on element fluxes and pools in waters

547

In typical natural environments subjected to wildfire, the mobilization of ash constituents from soil to surface waters is limited by water availability after a fire event, given the relatively arid climate conditions (i.e., Bodi et al., 2014). It has been shown that degrading soot (condensed volatiles) and char (solid organic residue) can move downward into the peat matrix (Certini, 2005; Preston and Schmidt, 2006; Flannigan et al., 2009; Olefeldt et al., 2013) thereby transferring inorganic and organic nutrients back into the soil (Ackley et al., 2021). Furthermore, in the majority of temperate and subtropical settings where wildfires occur, ash leachate products can be transported into groundwater (Hauer and Spencer, 1998; Bodi et al., 2014). These factors (water/moisture limitation and alternative pathways) certainly decrease the potential impact of wildfire on element delivery to the hydrological network. The permafrost nature of the Siberian tundra presents a stark contrast to these settings due to excessive precipitation (especially at the end of the active seasons) and impermeable permafrost layers that allow rapid, complete and (essentially surface) suprapermafrost flow-mediated transport of ash constituents to adjacent thermokarst lakes and draining rivers. In these conditions, the entire amount of elements present in leachable form in ash produced by burnt surface peat and ground vegetation can be mobilized into hydrological network.

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The mass balance calculations provided in section 3.3 (Fig. 5) present the most conservative scenario given that the ash produced from dwarf shrubs, grasses and forest floor litter burning is not included in the leachable stock and that the peat is burned only to a depth of 10 cm. Furthermore, a fire return interval of 56 years for permafrost palsa peatland, based on past decades of monitoring and paleo-reconstructions, may become much shorter in the coming years due to Pan-Arctic increases in summer heat waves and wildfire frequency at higher latitudes (Rocha et al., 2012). Thus, in Arctic tundra, wildfire is projected to increase 4- to 11-fold by 2100 due to increased temperature, evapotranspiration, ignition sources, and vegetation shifts (Abbott et al., 2016; Hu et al., 2010, 2015). Nevertheless, mass balance calculations
performed in the present study suggest that even in the case of 10-fold increase in wildfire frequency (and thus, a decreased fire return interval of 5-6 years), its impact on annual element export by tundra rivers will not exceed 10 to 20%. Such a small increase is within the uncertainties of river solute annual yield quantification. Moreover, even for the case of minimal territory coverage by lakes (4%), the ash dissolution in surface waters due to the 10-fold increase in wildfire frequency will not impact the concentration of most elements in thermokarst lake waters more than 10-20%, with the exception of likely Ca and PO₄. The later finding is consistent with well documented wildfire impact on P export in forested watersheds (Burke et al., 2005). At the same time, the effects of tundra wildfire may be drastically more important on a scale of a single burning event (i.e., Carignan et al., 2000), which can double the concentrations and export fluxes of most major and micro-nutrients for a particular year. In this regard, there is a clear need for further investigation of the significantly understudied reactivity of ash produced by burning of the tundra and forest-tundra zone organic litter horizons, deeper (> 10 cm) peat horizons, as well as individual species of grasses and dwarf shrubs.

At the same time, one has to keep in mind the restricted scale of experiments which were aimed solely to quantify to which degree the wildfire in permafrost peatland ecosystems can affect nutrient mobility from the burned soil/peat to the adjacent hydrological network. In particular, our approach ignores particulate emissions (PM2.5) present in smoke and soot, which would be mobilized in dryfall in the form of highly soluble compounds and thus enhance the overall impact of wildfire on surrounding aquatic ecosystems. Furthermore, on a broader scale, the main mechanisms for wildfire effects on biogeochemistry have to deal with altered albedo due to particles deposition from combustion, changes in the depth of the active layer, vegetation and soil moisture — all of which affect microbial activity and nutrient mineralization and mobility (cf., Jafarov et al., 2013).

## 5. Conclusions
Mosses, lichens, and peat represent the dominant organic substrates in permafrost peatlands subjected to wildfire. Interactions between ash derived from these materials and distilled and natural organic-rich thermokarst lake water were analyzed to assess element release into aqueous solutions. Results of these experiments demonstrated that the majority of solutes (>80-90%) were leached over the first 10 minute of reaction. This suggests virtually instantaneous release of ash constituents into aqueous solution. In accord with the first hypothesis of this study, moss ash is by far the most reactive substance. Moss ash was capable of increasing the pH from 8 to 10 while largely controlling concentrations of soluble highly mobile elements (anions and oxyanions [DIC, Cl, SO₄, B, Mo, Se and Sb], alkalis [Na, K, Rb and Cs], Mg, numerous micronutrients [B, Cu, Mn, Co, Ni, V, Mo and Se] and other trace elements [Al, Sb and U]. The enhanced mobility of oxyanions in moss ash might be due to the alkaline environment of aqueous solution promoting hydrolysis (Al and U<sup>VI</sup>) or oxyanion desorption from ash particles. Alternatively, lichen ash dissolution affected Si, P and the majority of low solubility hydrolysates [Fe, Y, Zr, Th and REEs] and atmosphere originated pollutants (Zn, Pb and Bi). This is likely due to preferential accumulation of lithogenic minerals, solid aerosols and/or formation of phosphate minerals in their tissues. Finally, peat ash released the highest amounts of Ca, Sr and Ba. This may be linked to enhanced dissolution of carbonate minerals given the rather acidic environments of solutions in which peat ash reacted.

The experimental results did not confirm the second hypothesis of this study, on the importance of tundra wildfire in element cycling in aquatic systems. Despite the very high reactivity of mosses, lichens and peat ash in surface waters, mass balance calculations (factoring in actual vegetation and peat stock of permafrost peatlands and typical regional wildfire frequency) demonstrate that the impact of burning on the element yield of rivers or the element storage of thermokarst lakes is negligibly small. Even a drastic (10-fold) increase in wildfire frequency cannot sizably (> 10-20%) modify macro- and micro-nutrient delivery to the hydrological network (except in the case of Ca and P for lakes). However, the estimations
performed in the present work represent a limited, on-side approach to much complex problem
of fire event in tundra ecosystems. This highlights the need for further experimental studies on
ash derived from forest litter, dwarf shrubs, grasses and perhaps deeper peat horizons in
permafrost peatlands.

628

629

630 Acknowledgements

The reported study was funded by RFBR and Yamalo-Nenets Autonomous Okrug,
project No 19-44-890013, the RFBR grant No 21-54-75001, and the TSU Development Program
“Priority-2030”. We thank two anonymous reviewers for their very thorough and constructive
reviews.

635

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<table><caption>**Table 1.** Element yield (mean ± s.d.) from peat, moss and lichen ash to aqueous solution (averaged for Milli-Q water and organic-rich lake water) over 167 h of reaction.</caption><thead><tr><th></th><th>Peat ash</th><th>Moss ash</th><th>Lichen ash</th></tr></thead><tbody><tr><td></td><td colspan="3" style="text-align: center;">mg/g<sup>-1</sup></td></tr><tr><td>DIC</td><td>0.20±0.20</td><td>4.71±4.90</td><td>0.24±0.25</td></tr><tr><td>DOC</td><td>1.68±1.68</td><td>4.00±4.22</td><td>2.41±2.61</td></tr><tr><td>SO<sub>4</sub><sup>2-</sup></td><td>5.69±0.10</td><td>11.26±0.11</td><td>2.48±0.12</td></tr><tr><td>Cl-</td><td>2.69±0.58</td><td>4.71±1.10</td><td>3.26±0.41</td></tr><tr><td></td><td colspan="3" style="text-align: center;">µg/g<sup>-1</sup></td></tr><tr><td>B</td><td>2.75±0.19</td><td>21.3±0.29</td><td>10.6±0.37</td></tr><tr><td>Na</td><td>274.8±79</td><td>1575±124</td><td>315±70.5</td></tr><tr><td>Mg</td><td>614±51</td><td>2330±86</td><td>96.3±15.4</td></tr><tr><td>Al</td><td>8.52±3.0</td><td>66.4±11.4</td><td>77.4±33.1</td></tr><tr><td>Si</td><td>115±19.7</td><td>490±39</td><td>718±41</td></tr><tr><td>P</td><td>599±93</td><td>814±59</td><td>3878±206</td></tr><tr><td>K</td><td>2429±502</td><td>20865±2280</td><td>7869±1450</td></tr><tr><td>Ca</td><td>1663±112</td><td>1566±86</td><td>95.6±22.9</td></tr><tr><td>Ti</td><td>0.10±0.049</td><td>0.21±0.11</td><td>3.77±1.73</td></tr><tr><td>V</td><td>0.13±0.013</td><td>2.35±0.18</td><td>0.15±0.038</td></tr><tr><td>Cr</td><td>0.05±0.035</td><td>0.44±0.0069</td><td>0.46±0.15</td></tr><tr><td>Mn</td><td>0.82±0.11</td><td>27.9±4.6</td><td>2.91±0.49</td></tr><tr><td>Fe</td><td>2.81±1.69</td><td>3.52±3.04</td><td>62.3±28.8</td></tr><tr><td>Co</td><td>0.044±0.0051</td><td>0.22±0.049</td><td>0.051±0.010</td></tr><tr><td>Ni</td><td>0.059±0.0080</td><td>1.21±0.29</td><td>0.21±0.11</td></tr><tr><td>Cu</td><td>0.040±0.0087</td><td>0.19±0.031</td><td>0.22±0.11</td></tr><tr><td>Zn</td><td>1.46±0.14</td><td>0.82±0.47</td><td>2.74±0.66</td></tr><tr><td>As</td><td>0.26±0.022</td><td>0.94±0.014</td><td>0.89±0.073</td></tr><tr><td>Se</td><td>0.10±0.0087</td><td>0.44±0.027</td><td>0.11±0.010</td></tr><tr><td>Rb</td><td>0.69±0.066</td><td>63.0±2.1</td><td>18.0±0.81</td></tr><tr><td>Sr</td><td>14.3±0.53</td><td>2.71±0.08</td><td>0.28±0.067</td></tr><tr><td>Y</td><td>0.0023±0.00042</td><td>0.0023±0.0010</td><td>0.010±0.0063</td></tr><tr><td>Zr</td><td>0.0082±0.0037</td><td>0.0087±0.0047</td><td>0.16±0.056</td></tr><tr><td>Mo</td><td>0.094±0.014</td><td>0.55±0.035</td><td>0.18±0.010</td></tr><tr><td>Cd</td><td>0.0040±0.00038</td><td>0.0030±0.0014</td><td>0.0069±0.0026</td></tr><tr><td>Sb</td><td>0.063±0.0037</td><td>0.70±0.048</td><td>0.32±0.016</td></tr><tr><td>Cs</td><td>0.021±0.0022</td><td>0.96±0.047</td><td>0.24±0.014</td></tr><tr><td>Ba</td><td>6.17±0.39</td><td>2.38±0.094</td><td>0.38±0.085</td></tr><tr><td>La</td><td>0.0012±0.00025</td><td>0.0026±0.0018</td><td>0.015±0.0077</td></tr><tr><td>Ce</td><td>0.0021±0.00031</td><td>0.0039±0.0039</td><td>0.028±0.012</td></tr><tr><td>Pr</td><td>0.00027±0.000034</td><td>0.00041±0.00041</td><td>0.0035±0.0017</td></tr><tr><td>Nd</td><td>0.0011±0.00018</td><td>0.0018±0.0014</td><td>0.013±0.0068</td></tr><tr><td>Sm</td><td>0.00041±0.000087</td><td>0.0026±0.0037</td><td>0.0033±0.0018</td></tr><tr><td>Eu</td><td>0.00077±0.000037</td><td>0.00035±0.000075</td><td>0.00062±0.00031</td></tr><tr><td>Gd</td><td>0.00025±0.000080</td><td>0.00030±0.00027</td><td>0.0023±0.0013</td></tr><tr><td>Tb</td><td>0.000032±0.0000081</td><td>0.000044±0.000031</td><td>0.00032±0.00017</td></tr></tbody></table>
<table><tr><td>Dy</td><td>0.00017±0.000024</td><td>0.00024±0.00018</td><td>0.0018±0.00090</td></tr><tr><td>Ho</td><td>0.000036±0.0000044</td><td>0.000048±0.000032</td><td>0.00036±0.00020</td></tr><tr><td>Er</td><td>0.00011±0.000023</td><td>0.00014±0.000091</td><td>0.0010±0.00058</td></tr><tr><td>Tm</td><td>0.000019±0.0000037</td><td>0.000020±0.000013</td><td>0.00014±0.000082</td></tr><tr><td>Yb</td><td>0.00012±0.000021</td><td>0.00013±0.000087</td><td>0.00088±0.00053</td></tr><tr><td>Lu</td><td>0.000020±0.0000036</td><td>0.000022±0.000015</td><td>0.00013±0.000078</td></tr><tr><td>W</td><td>0.0026±0.0017</td><td>0.23±0.023</td><td>0.10±0.014</td></tr><tr><td>Tl</td><td>0.0056±0.0047</td><td>0.0071±0.0044</td><td>0.0047±0.00027</td></tr><tr><td>Pb</td><td>0.015±0.0036</td><td>0.046±0.028</td><td>0.27±0.094</td></tr><tr><td>Bi</td><td>0.0013±0.00069</td><td>0.0016±0.00060</td><td>0.010±0.0047</td></tr><tr><td>Th</td><td>0.00076±0.00091</td><td>0.0011±0.00081</td><td>0.035±0.0032</td></tr><tr><td>U</td><td>0.00013±0.000010</td><td>0.013±0.0022</td><td>0.0012±0.00070</td></tr></table>

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Peat Moss Lichen
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<figure><img src="image_11.png" /><figcaption>**Fig. 1.** The pH and SO₄²⁻ (left panels, A, C, E, G) and Specific Conductivity and Cl⁻ (right panels, B, D, F, H) evolution as a function of elapsed time in ash dissolution experiments in the Milli-Q water and lake water. The symbols are mean values of three independent experiments, whereas the error bars represent s.d. uncertainty of triplicates (unless not shown because they are within the symbol size).</figcaption></figure>

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<figure><img src="image_12.png" /><figcaption>**Figure 2.** Example of 1st group element evolution in the course of ash leaching in MilliQ water (left panels) and lake water (right panels): Si (A), P (B), Mn (C), Co (D), As (E), Mo (F), Sb (G) and Rb (H). The symbols are mean values of three independent experiments, whereas the error bars represent s.d. uncertainty of triplicates (unless not shown because they are within the symbol size).</figcaption></figure>
Figure 3. Example of 2nd group element evolution (decreasing concentration with time) in the course of ash leaching in Milli-Q water and lake water: Cu, Zn, Cd, DOC and SUVA<sub>254</sub>. The symbols are mean values of three independent experiments, whereas the error bars represent s.d. uncertainty of triplicates (unless not shown because they are within the symbol size).

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■ Peat/MQ ■ Peat/Lake water ■ Moss/MQ ■ Moss/Lake water ■ Lichen/MQ ■ Lichen/Lake water

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**Figure 4.** Elementary yields (normalized to the mass of ash) over the first 2 days and subsequent 7 days of reaction with peat, moss and lichen ash. The columns and error bars represent the mean values and s.d. of three independent experiments for each solid and liquid substrate. The red circle indicates significant differences (at p < 0.05) between Milli-Q and lake water.

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**Figure 4**, continued.

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**Figure 5.** Histogram of ratio (in %) of annual element leaching from burned vegetation in the palsa watershed of permafrost peatlands assessed in this study (Table S5) to annual elementary yield by small and medium size rivers in permafrost impacted portions of the WSL (from Pokrovsky et al., 2020).
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Experiment: organic material burned

peat

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moss

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lichen

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Tundra wildfire

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Negligible impact

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Discontinuous

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Ash

leached into lakes, rivers

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