Eur J Forest Res (2012) 131:1245–1253 DOI 10.1007/s10342-012-0595-8
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ORIGINAL PAPER
Carbon loads, forms and sequestration potential within ash deposits produced by wildfire: new insights from the 2009 ‘Black Saturday’ fires, Australia Cristina Santı´n • Stefan H. Doerr • Richard A. Shakesby Rob Bryant • Gary J. Sheridan • Patrick N. J. Lane • Hugh G. Smith • Tina L. Bell
•
Received: 26 July 2011 / Revised: 20 December 2011 / Accepted: 24 January 2012 / Published online: 10 February 2012 Ó Springer-Verlag 2012
Abstract Forest fires release substantial amounts of carbon (C). Much of it is emitted to the atmosphere, but some is deposited within ash on the ground. Little is known about amount and types of C deposited in ash. Here, we quantify total C, and total inorganic, water-soluble and particulate organic fractions deposited in ash during the catastrophic 2009 ‘Black Saturday’ wildfires in Australia. These fires coincided with the highest air temperatures and lowest humidity ever recorded in the local area, which, combined with high fuel loads of mostly long unburnt eucalypt forests, generated extreme burning conditions. In three mixed-species eucalypt forest sites sampled, the canopy,
This article originates from the Session ‘Ash in the Environment’ held at the General Assembly 2011 of the European Geosciences Union in Vienna, 3–8 April 2011.
understorey and litter fuels were almost completely consumed, resulting in substantial ash deposition (mean, 81.9 t ha-1), with 5.9 t ha-1 of C being transferred from vegetation to the forest floor. In five temperate rainforest sites sampled, the canopy was not burnt and ash deposition was lower (mean, 48.3 t ha-1) than in the mixed-species eucalypt forest, but overall their higher C content resulted in higher C deposition (8.1 t ha-1). In all cases, most C contained in ash was organic and its pyrogenic nature infers increased resistance to degradation. Pyrogenic C is viewed by many as an important C sink, which could contribute to long-term C sequestration when incorporated into soils or sediments. Our results highlight the potential importance of the pyrogenic C pool in freshly deposited ash and, therefore, the necessity of a systematic and detailed analysis of ash deposition and C forms in ash to improve our understanding of C fluxes by forest fires.
Communicated by A. Merino.
Keywords Mixed-species eucalypt forest Temperate rainforest Black carbon Charcoal Pyrogenic carbon
C. Santı´n S. H. Doerr (&) R. A. Shakesby College of Science, Swansea University, Swansea, UK e-mail:
[email protected]
Introduction
R. Bryant College of Engineering, Swansea University, Swansea, UK G. J. Sheridan P. N. J. Lane Department of Forest and Ecosystem Science, The University of Melbourne, Melbourne, VIC, Australia H. G. Smith School of Geography, Earth and Environmental Sciences, Plymouth University, Plymouth, UK T. L. Bell Faculty of Agriculture, Food and Natural Resources, University of Sydney, Sydney, NSW, Australia
Wildfires can transfer carbon (C) stored in biomass, necromass, litter and soil to the atmosphere, mainly as CO2. They contribute significantly to global CO2 emissions, amounting to a third of current annual emissions from fossil fuel burning and from industry (Flannigan et al. 2009). Fires (excluding those associated with deforestation), however, are usually thought to be long-term, ‘net zero C emission events’, since release of C during burning is expected to be balanced with uptake by regenerating vegetation (Bowman et al. 2009). Notwithstanding this basic concept, the role of fire in the C cycle is actually more complex. For instance,
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fire may indirectly affect the C cycle by causing changes in vegetation cover or ecological succession patterns, thereby modifying C stocks in the biomass (Bond-Lamberty et al. 2007; Turner et al. 1995). It may also alter soil C stocks by enhancing the redistribution of C through post-fire erosion and off-site burial of soil and organic layers (Shakesby and Doerr 2006), changing soil microbial and faunal activity (Certini 2005) and altering soil organic matter quantity and composition (Gonza´lez-Pe´rez et al. 2004). A common outcome of wildfires is the production of pyrogenic C. As a result of incomplete combustion, some of the biomass is transformed into pyrogenic organic matter (i.e. charred organic materials, charcoal, black carbon) (Baldock and Smernik 2002). These pyrogenic organic materials are mainly comprised of condensed refractory compounds, which are highly resilient to degradation and, therefore, likely to remain stored in soils or in terrestrial and aquatic sediments for a much longer period than is needed for full biomass recovery (Rodionov et al. 2010; Woolf et al. 2010). After a fire-vegetation regrowth cycle, pyrogenic C, therefore, will represent an added C stock and, thus, could be considered as a true type of C sink (Lehmann et al. 2008). The majority of pyrogenic organic compounds produced during wildfires are deposited in ash (Forbes et al. 2006). Post-fire ash deposits are defined as the solid residue remaining on-site after biomass burning and consist of a range of residual mineral materials and organic compounds affected by different grades of charring (definition modified from Scott 2010). To date, the role of ash has rarely been considered in studies examining C fluxes from wildfires, and the few data available have been obtained mainly from relatively low intensity prescribed fires (Forbes et al. 2006). One reason for this knowledge gap may be the rapid rate of ash redistribution within, and removal from burnt sites by wind and water erosion (Cerda` and Doerr 2008), which often occurs before the commencement of post-fire field studies, so that the potential importance of this source of C deposit may not always be obvious. Moreover, where ash has been considered, its C component has commonly been viewed as a unique category of pyrogenic material (see review by Forbes et al. 2006). This, however, is an oversimplification in the context of the longer-term effect of fires on C budgets. Different types of C compounds, with different origins and degrees of resistance to degradation, make up the C pool in ash. First, some can be present as inorganic C (e.g. calcium carbonate) produced during the combustion process (Liodakis et al. 2005). Second, pyrogenic organic compounds in ash may be derived from different types of vegetation undergoing different degrees of charring. Both factors condition their chemical composition and recalcitrance (Ascough
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et al. 2011; Knicker 2007). Third, variation in the physical properties of ash (e.g. particle size, solubility, texture), and in particular, of its organic components, may influence its mobilization and degradation in the environment and, hence, its role as a potential C sink. To address the knowledge gaps identified above, this study quantifies: (1) ash deposition, (2) the amount of C contained in the ash layer, and (3) the different forms of C in the ash (total inorganic, and total organic, watersoluble and particulate organic fractions) for forested sites that were burnt during the catastrophic 2009 ‘Black Saturday’ fires in Victoria, Australia. Given that these fires were the most extreme on record in this region, our results provide valuable insights into the potential impact on pyrogenic C production of future extreme fires, which are expected to become more common under predicted global warming scenarios (Flannigan et al. 2009; Liu et al. 2010).
Study sites and methods The ‘Black Saturday’ fires and sampling design The ‘Black Saturday’ wildfires in early February 2009 affected some 450,000 ha of eucalypt forest in Victoria, Australia and caused the tragic loss of 173 lives (Royal Commission 2009). Weather conditions were extreme, with dry northerly winds gusting up to 100 km h-1 coinciding with the highest recorded air temperatures and lowest humidity in this region, exceeding 45°C in many places (National Climate Centre 2009). These conditions, combined with high biomass associated with mature temperate eucalypt forests, unburnt by a major fire since 1939 (understorey fuel load including litter: 25–40 t ha-1), very low fuel moisture conditions (3–4%) and steeply dissected topography, generated burning conditions of hitherto unprecedented severity (McCaw et al. 2009). Average fireline intensity is estimated to have exceeded 70,000–80,000 kW m-1, which is substantially higher than any previously reported in Australia (Royal Commission 2009). Following extensive field reconnaissance of the Kilmore-Murrindini fire complex in the eastern region affected by wildfire and using fire severity maps (available from Victorian Department for Sustainability and the Environment), three field sites within 3 km of Marysville were selected, including Acheron (Ah), Mount Gordon (Gh) and Marysville Beauty Spot (Mh). These sites displayed the most severe burn conditions, classified as ‘Extreme fire severity’ based on the amount of canopy burnt and ground observations using the degree of combustion of the forest floor biomass (Chafer et al. 2004). This class involved complete consumption of leaf litter, green vegetation and
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Fig. 1 Mixed-species eucalypt forest at Marysville Beauty Spot (Mh). The extreme fire severity is apparent from the complete elimination of leaf litter, shrubs and smaller diameter (\10 mm) branches [cf. fire severity classification of Chafer et al. (2004)] and a thick layer of ash
woody stems \10 mm in diameter (Fig. 1). These sites, which were last burnt by wildfire in 1939,1 were representative of the most common forest type in the region (mixed-species eucalypt forest) and had slope angles of less than 5°. Additional sampling was carried out in nearby temperate rainforest (Rf), at five sites located approximately 15 km SSW of Marysville and within a radius of 5 km of each other. The temperate rainforest sites were characterized by tall ([30 m high) Mountain Ash (Eucalyptus regnans F.Muell.) trees with a dense understorey containing tree ferns (Dicksonia spp.) and a variety of woody shrubs. At these sites, fire severity was classified as ‘Moderate to high’ (Chafer et al. 2004). Here, the tree canopy was largely unaffected by wildfire; however, the understorey vegetation and the leaf litter layer were fully consumed (Fig. 2). Grid references (UTM) for all sites are provided in Table 1. Sampling was undertaken in autumn (April) 2009 following some light post-fire rainfall, but before intense rainfall events of winter had caused any significant erosion of ash. For each site in the mixed-species eucalypt forest (Ah, Gh, Mh), four parallel 20 m transects were laid out 5 m apart and sampled at 5 m intervals. At each of the five temperate rainforest sites (Rf), one 20 m transect was sampled at 1 m intervals. At each sampling point, a 30 9 30 cm square was sampled. Ash depths were measured in the centre of four 15 9 15 cm sub-squares and the entire ash layer was carefully collected. The cohesive nature of the typically moist ash layer contrasted with the 1
Sites Ah and Mh experienced a prescribed fuel reduction burn in 1988 and 1982 respectively. Given the long recovery period ([20 years) until the 2009 fires, fuel loads can be considered to have fully recovered.
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Fig. 2 Burnt temperate rainforest (Rf). The fine understorey fuel and vegetation layer were consumed. The canopy of E. regnans trees remained largely unaffected and this site was classified as having been affected by moderate to high burn severity (cf. Chafer et al. 2004). Post-fire litter fall and downed wood has provided some ground cover
dry and highly water repellent underlying mineral soil (Doerr et al. 2010) facilitating complete and uncontaminated sampling of ash. Ash load per unit area was determined based on the air-dry weight of the ash material collected from each sampling area. A long unburnt site (Pc; last burn in 1939) was selected approximately 6 km WSW of Marysville (UTM: 55H 0378605, 5843222) to provide an estimate of the fine fuel load associated with the understorey vegetation (\10 mm diameter) and leaf litter layer. This site was selected because it was as similar as possible to the main burnt sites (Ah, Gh, Mh) in terms of forest type, time elapsed since the previous fire and underlying geology. Laboratory analysis Following air-drying, ash samples from each transect were bulked into single samples resulting in four composite samples for each of the mixed-species eucalypt forest sites (Ah, Gh, Mh) and five for the temperate rainforest (Rf) sites. Pooled ash samples were analysed (see schematic in Fig. 3) to determine: total carbon (TC), total inorganic carbon (TIC), total organic carbon (TOC), water-soluble organic carbon (WSOC) and water-extractable particulate organic carbon (WE-POC) using the procedures described below. TC was determined from ground subsamples by dry combustion at 1,050°C (Primacs TC Analyser, Skalar UK). TIC content was determined from a separate subsample, which was heated in air for 4 h at 450°C (Cambardella et al. 2001) to eliminate the organic C by combustion, and then analysed to determine the TC remaining. Following
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Table 1 Characteristics of ash and estimated carbon load for all study transects Site and UTM coord.
Transect
Deptha (cm)
Bulk densitya (g cm-3)
Ash load (g m-2 [t ha-1])
TC load (g C m-2 [t C ha-1])
TOC load (g C m-2 [t C ha-1])
TIC load (g C m-2 [t C ha-1])
Mixed-species eucalypt forest Ah
1
1.4 (±1.6)
0.5 (±0.3)
4793
370
359
11
55H
2
2.3 (±0.5)
0.3 (±0.1)
8196
611
602
8
0383323
3
1.8 (±0.3)
0.4 (±0.1)
7013
598
593
6
5842264
4
2.0 (±0.7)
0.4 (±0.3)
7807
601
596
5
1.9
0.4
6952 [69.5]
545 [5.4]
538 [5.4]
7 [0.1]
Mean Gh 55H
1
1.0 (±0.3)
0.4 (± 0.3)
4721
400
394
6
2
1.5 (±0.9)
0.5 (±0.0)
7075
680
677
3
0386920
3
1.2 (±0.8)
0.6 (±0.1)
5681
440
428
11
5845121
4
2.5 (±2.0)
0.4 (±0.3)
6733
417
417
No data
1.7
0.5
6053 [60.5]
484 [4.8]
479 [4.8]
7 [0.1]
Mean Mh
1
2.3 (±0.9)
0.5 (±0.1)
10603
651
646
5
55H
2
2.9 (±1.4)
0.4 (±0.1)
10256
649
633
16
0391031 5847757
3 4
3.8 (±1.9) 1.7 (±0.9)
0.4 (±0.1) 0.4 (±0.2)
15195 10172
1000 694
968 673
32 20
Mean
2.7
0.4
11557 [115.6]
748 [7.5]
730 [7.3]
18 [0.2]
Mean Ah, Gh, Mh
2.0
0.4
8187 [81.9]
593 [5.9]
582 [5.8]
11 [0.1]
1
3.0 (±3.2)
0.2
5277
970
964
6
2
2.8 (±2.5)
0.2
4612
856
856
No data
3
1.4 (±1.1)
0.3
4547
730
730
No data
4
1.2 (±1.3)
0.4
4665
550
550
No data
5
1.5 (±1.4)
0.4
5492
927
913
13
2.0
0.3
4829 [48.3]
806 [8.1]
802 [8.0]
10 [0.1]
Temperature rainforest Rfb
Mean
TC total carbon, TOC total organic carbon, TIC total inorganic carbon a
Mean values (± standard deviation)
b
UTM coordinates: Rb1: 55H 0387165, 5833175; Rb2: 55H 0385490, 5835570; R3: 55H 0385734, 5836679; R4: 55H 0386021, 5836508; R5: 55H 0385220, 5836738
correction to account for the weight loss that occurred when eliminating the organic C, this value was assumed to represent TIC. TOC content was calculated as the difference between TC and TIC. WSOC was quantified according to Ghani et al. (2003). Ash samples (2 g) were weighed into 50-ml polypropylene centrifuge tubes and shaken with double-distilled water (40 ml) for 30 min at 20°C (±1°C) and centrifuged for 20 min at 3,500 rpm. The supernatants were removed and vacuum-filtered through 0.45-lm cellulose nitrate membranes. The filtrates were analysed for total dissolved C by high-temperature catalytic oxidation to CO2 at 680°C and subsequent quantification by a non-dispersive infrared gas analyzer (Shimadzu TOC-V CPH). The inorganic dissolved C was determined following addition of hydrochloric acid to a replicate sample and quantifying the CO2 generated. This value was subtracted from the total dissolved C to give WSOC. The
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particulate material ([0.45 lm) captured by the cellulose nitrate membranes represents the WE-POC (Fig. 3) and was freeze-dried, weighed, ground and its TC content determined (Primacs TC Analyser). Due to its organic nature, the TC content was considered to represent TOC and was thus designated as WE-POC. For all analyses, the difference between replicates of individual measurements was less than 5%. The weight of the fine fuel load (leaf litter and small diameter branches of \10 mm) was estimated from an airdried subset of the samples taken from the long unburnt site (Pc). Given that all of the fine fraction of the understorey fuel had been consumed at the burnt sites, the understorey fuel load sampled at the long unburnt site (Pc) was considered to be representative of the fine fuel load at the burnt sites Ah, Gh and Mh. The resulting value is an underestimation of the total fuel load as data enabling the
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Ash sample bulked from transect sites
Combustion (4 h at 450 ºC)
Dispersion in water Centrifugation
Solid carbon analysis
Solid
Supernatant liquid Filtration at 0.45 µm
Total Inorganic Carbon (TIC)
0.45 µm
Total Carbon Solid Carbon Analysis Total Organic Carbon (TOC)
Acidification
Catalytic oxidation and CO2 analysis
Water Extractable Particulate Carbon (WE-POC) Total soluble carbon
Soluble inorganic carbon
Water Soluble Organic Carbon (WSOC)
Fig. 3 Analytical procedures for extraction and quantification of the different forms of carbon in ash (TC, TIC, TOC, WE-POC and WSOC)
quantification of the contribution from the canopy, bark and coarse woody debris were not available.
Results and discussion Ash deposition Mean depth and bulk density of ash (calculated from ash depth and mass per unit area) for each transect and site is given in Table 1. Ash depth varied considerably between sampling points, ranging from no ash to more than 8 cm (mean 2.0 cm) for the mixed-species eucalypt forest and from 0.3 to 10 cm (mean 2.0 cm) for the temperate rainforest. The greatest average depth of ash was recorded at Mh (2.7 cm). Although all three mixed-species eucalypt forest sites were subject to extreme burn severity, the greater depth of ash at site Mh could indicate that this site was at the highly destroyed end of this category (Fig. 1) and/or the fuel load here was higher here than at the other two sites. Mean bulk density of ash ranged from 0.3 to 0.6 g cm-3 for mixed-species eucalypt forest and from 0.2 to 0.4 g cm-3 for temperate rainforest (Table 1). The mean ash load for the mixed-species eucalypt forest sites ranged from 60.5 to 115.6 t ha-1 (mean 81.9 t ha-1), which was
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relatively high compared to an average of 48.3 t ha-1 for temperate rainforest. At the temperate rainforest sites, fire burnt the understorey vegetation but did not reach the canopy, which could explain the resultant smaller ash load for this forest type. Relatively few data have been documented in relation to ash deposition after fire. Bulk density of ash reported after severe wildfire in conifer forests ranges from 0.18 to 0.62 g cm-3 (Cerda` and Doerr 2008; Goforth et al. 2005; Woods and Balfour 2008). Our values fall within this range, which suggests that: (1) prior to sampling, no substantial compaction of the ash layer occurred because of post-fire rainfall; and (2) during sampling, there was little, if any, contamination of the ash by the underlying mineral soil. There are no values available for eucalypt forests, but ash depths and loads reported here are comparable with those for conifer forest in Montana (mean depth and load, 1.9 cm and 77.9 t ha-1; Woods and Balfour 2008), California (0.6–0.8 cm and 14.4–27.0 t ha-1; Goforth et al. 2005) and Spain (3.6 cm and 151.2 t ha-1; Cerda` and Doerr 2008). However, ash loads were substantially higher than expected from consumption of dry, understorey fine fuel from the unburnt mixed-species eucalypt forest site, Pc (28.6 t ha-1). This does not mean that this site is unrepresentative of this forest type, but more likely that the difference in measured and anticipated ash load does not take into account the contribution of ash from combustion of the canopy, bark on standing trees and coarse woody debris, for which no reliable information exists from the area burnt in 2009. Coarse woody debris, for example, can be a major fuel component as shown by Tolhurst et al. (1992) for a long unburnt (51 years) dry sclerophyll eucalypt forest site in eastern Victoria. They estimated fine fuel load to be 17.1 t ha-1, whereas the coarse fuel load (dead woody material [25 mm on the ground) was as high as 105.1 t ha-1. C loads and forms in ash For the mixed-species eucalypt forest (Ah, Mh, Gh), estimated TC loads in ash ranged from 3.7 to 10.0 t TC ha-1, with a mean of 5.9 t TC ha-1 (593 g TC m-2; Table 1). For the temperate rainforest (Rf), TC loads were higher despite having lower ash loads than the mixed-species eucalypt forest (8.1 t C ha-1). Irrespectively of TC load, only a small fraction of C was attributable to TIC (0.1–0.2 t C ha-1; 3–32 g C m-2). These values are comparable with the lower figures of inorganic C deposition in ash reported after high-severity wildfires in conifer forest and oak woodland in California (0.1–2.7 t C ha-1, Goforth et al. 2005). Absolute amounts and proportions of different forms of C in ash are given in Table 2. Organic C clearly dominated
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Table 2 Different carbon forms in ash for all study transects Site
Transect
TC (mg C g-1)
TOC (mg C g-1)
TIC (mg C g-1)
WE-POC (mg C g-1)
WSOC (lg C g-1)
Mixed-species eucalypt forest Ah
1
77.2
75.0
2.2
13.5
129.3
2
74.5
73.5
1.0
18.1
136.1
3
85.3
84.5
0.8
11.6
170.4
4
77.0
76.4
0.6
38.8
Mean (% of TC) Gh
78
1 (1)
21 (26)
157.6 148 (0.2)
1
84.8
83.5
1.3
13.9
327.2
2
96.1
95.7
0.4
7.4
371.2
3
77.4
75.4
2.0
2.8
384.3
4
62.0
62.0
n.d.
5.7
Mean (% of TC) Mh
77 (99)
80
79 (99)
1 (1)
8 (9)
340.0 356 (0.5)
1
61.4
60.9
0.5
2.8
261.1
2
63.3
61.7
1.6
3.7
221.9
3
65.8
63.7
2.1
2.5
181.3
4
68.2
66.2
2.0
6.4
271.8
Mean (% of TC)
65
63 (98)
2 (2)
4 (6)
234 (0.4)
Mean Ah, Gh, Mh (% of TC)
74
73 (98)
1 (2)
11 (14)
246 (0.3)
Temperature rainforest Rf
Mean (% of TC)
1
183.8
182.6
1.2
78.1
329.0
2
185.5
185.5
n.d.
34.7
377.8
3
160.5
160.5
n.d.
55.9
406.4
4 5
117.8 168.7
117.8 166.3
n.d. 2.4
14.9 33.0
225.1 274.3
163
163 (99)
2 (1)
43 (26)
323 (0.2)
TC total carbon, TOC total organic carbon, TIC total inorganic carbon, WEPOC water-extractable particulate organic carbon, WSOC watersoluble organic carbon n.d. not detected
in all samples accounting for over 97 to nearly 100% of the TC, which explains the similarity of estimated loads of TC and TOC. Deposited ash from the mixed-species eucalypt forest sites (Ah, Gh, Mh) had an average TOC content of 73 mg C g-1, whereas ash from temperate rainforest was relatively richer in organic C (163 mg C g-1). This resulted in higher loads of TC. The greater organic content of ash from temperate rainforest sites may be due to incomplete combustion of fine fuel given that fire intensities were moderate to high and fuel moisture levels may have been somewhat higher here compared to the mixed-species sites, particularly for the layer of litter closest to the mineral soil. The combustion of fine fuel depends on, amongst other factors, water content (Matthews et al. 2007). The inorganic C contribution to TC was very low in all ash samples (\3%), which is commonly reported for ‘dark ´ beda et al. 2009) although this type of ash is supash’ (U posedly characteristic of low-severity fires and usually has a higher TOC content than found in our samples (e.g. Bodı´ et al. 2010).
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Estimates of C deposition from analysis of ash need to reflect that it is non-uniformly distributed and present in various chemical forms. Although the TIC contribution is minor in our samples (Table 2), carbonates and bicarbonates have been reported as substantial constituents in ash produced by other fuel types and/or generated under other ´ beda et al. burning conditions (Gabet and Bookter 2011; U 2009; Ulery et al. 1993). For example, Goforth et al. (2005) found that inorganic C, present as calcium carbonate, accounted for up to 10% of sample weight for ‘white ash’ from a severely burnt mixed-species conifer forest. The assumption that TOC can be determined simply from TC may, therefore, not always hold true (see Forbes et al. 2006 and references therein). In addition, the organic fraction of ash is a mixture of organic compounds derived from different types of vegetation and will have been affected by different degrees of charring, so that diverse chemical and physical characteristics can be expected in this fraction. The water-extractable particulate (WE-POC) and water-soluble (WSOC)
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organic C contents in ash are shown as both absolute values and relative contributions to TC in Table 2. In our study, WE-POC accounts for 6–26% of the TC in ash, which is quantitatively an important contribution, especially given that particulate organic matter (POM) is of low density and not bound to minerals. Consequently, in burnt surface soil, POM is highly susceptible to post-fire erosion by overland flow and likely to be preferentially redistributed relative to other organic components on hillslopes (Rumpel et al. 2006) and transported downstream in the channel network (Smith et al. 2011). Furthermore, while smaller organic components may be attached to minerals in soils and sediments and thereby stabilized and protected from degradation by various physical mechanisms (e.g. occlusion, sorption), the stability of free POM is mainly controlled by its inherent molecular-level resistance (Marschner et al. 2008). Due to its pyrogenic origin, such ash-derived POM is expected to have both a high chemical recalcitrance and high resistance to biodegradation (Keiluweit et al. 2010), making it one of the most stable forms of POM (Helfrich et al. 2006). The other fraction of C quantified here is water-soluble organic C (WSOC; Table 2), which is an important potential contributor to the dissolved organic C (DOC) in soils (Rennert et al. 2007). In our samples, WSOC concentrations ranged between 129.3 and 406.4 lg C g-1, which is higher than those recorded for laboratory-produced ash from wetland plants (48.6–232.3 lg C g-1; Zhao et al. 2010). These concentrations represent on average only 0.2–0.4% of the TC in ash. Despite this, DOC from pyrogenic materials (i.e. ash) should not be neglected since it is likely to contain small-sized polyaromatic hydrocarbons (Dittmar 2008). These are well-known pollutants in waters (European Commission 2000) and their production by fires could threaten water quality (Smith et al. 2011; Vila-Escale´ et al. 2007). In addition, bioavailability of DOC is crucial in cycling of organic matter in both aquatic and terrestrial systems, because most microbial transformation processes require a soluble phase (Marschner and Kalbitz 2003). Furthermore, bioavailability of DOC is often negatively correlated with the content of aromatic compounds (Marschner et al. 2008). It would, therefore, be expected that WSOC derived from ash contributes mainly to the ‘recalcitrant’ pool of DOC (Dittmar and Paeng 2009). From a physical point of view, our results for dissolved and particulate organic fractions suggest that not all the C in ash will exhibit the same behaviour. For instance, it has been suggested that POC and DOC account for most of the C eroded from soils and are, therefore, redistributed in the landscape or accumulate in depositional sinks (Jacinthe et al. 2001). In addition, important chemical differences are also expected within the pyrogenic fraction contained in
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ash. Recently, Ascough et al. (2011) identified a wide range in chemical composition of naturally produced and laboratory-generated charcoals. Laboratory burning typically produces distinct and more homogeneous pyrogenic materials for each specific experimental setting (Keiluweit et al. 2010). Naturally produced pyrogenic materials, on the other hand, result from charring of heterogeneous fuel materials under variable fire conditions and, therefore, represent a more complex range of pyrogenic materials (Knicker 2007; Knicker et al. 2008). Chemical recalcitrance is one of the main parameters determining the mean residence time of organic matter in the environment (Cheng et al. 2008; Marschner et al. 2008) and is, therefore, a key factor in its role as a C sink. The chemical composition of the organic component of ash is clearly a topic that warrants further investigation. Implications for C sequestration In a separate study conducted using the same sites examined here, Doerr et al. (2010) found that the burn severity of soil during the 2009 fires was less than anticipated given the extreme fire severity indicated by the extensive degree of vegetation destruction. They also found that the seedbank in the surface soil was only partially affected so that complete regeneration of biomass would be expected for most sites on a decadal scale. Over such a timescale, this particular fire could, therefore, be conservatively viewed as a ‘zero C emission’ event (Bowman et al. 2009). If, however, the C stored in ash is also taken into account, the data presented here demonstrate that wildfires may actually lead to significant C sequestration (once the forest biomass has fully restocked). The ash loads found here represent C deposition of 4.8–8.1 t C ha-1 on the forest floor, because part of the C previously stored in biomass has been transferred to the pool of C on the ground. Site visits in July 2009 revealed that much of the ash had subsequently been redistributed and that some of it had accumulated on footslopes and in depressions. Some ash is likely to have entered the fluvial system and can be expected to form part of river, lake and, ultimately, marine deposits. In any case, it would be protected to some degree from further erosion and consumption in future fires that might occur once a sufficient fuel load has been re-established. Decomposition rates of C buried at depositional sites would be reduced since they are usually characterized by low microbial activity, limited oxygen availability and a lack of accessible labile C sources (Knicker 2007; Kuzyakov et al. 2009). In addition, most C in our ash samples is organic (Table 2) and pyrogenic organic compounds are expected to be more recalcitrant and to have a much longer half-life than unburnt biomass (Marschner et al. 2008). These factors acting in combination could be expected to promote the
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long-term protection of C deposited in ash during this fire event, hence representing a true sequestration of C within the soil profile and/or in terrestrial or aquatic sediments. In fact, the industrial-scale conversion of biomass into recalcitrant C forms (i.e. biochar) is currently being considered as one of the most promising strategies for C sequestration in a safe and potentially useful solid form (Macı´as and Camps-Arbestain 2010; Woolf et al. 2010).
Conclusions Despite the extreme nature of the 2009 ‘Black Saturday’ wildfires, ash depth and bulk density and the resultant ash load were within the range of those few reported elsewhere in association with severe wildfires. For the forest types examined here, it was found that a substantial amount of C was present in ash. Wildfire led to an estimated deposition of 4.8–8.1 t C ha-1 in ash, representing transfer of part of the C pool held in unburnt biomass and litter prior to the fire to the forest floor. It is well established that pyrogenic C can be very resistant to degradation; however, the potential significance of substantial amounts of C being deposited in ash is only now being realized. The long-term fate of ash and the C contained therein is currently only poorly understood and requires further investigation. Notwithstanding this uncertainty, it is feasible that some types of wildfire could be contributing to natural C sequestration similar to what it is hoped that biochar production can achieve artificially through industrial-scale C sequestration. Acknowledgments C. Santı´n is grateful to the Alfonso Martin Escudero Foundation for a postdoctoral fellowship. This project was supported by a United Kingdom Natural Environment Research Council Urgency Grant (NE/F00131X/1). The authors thank C. Sherwin for providing fire history information and two anonymous reviewers for helpful comments and suggestions.
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