• Open Access

Aboveground plant biomass, carbon, and nitrogen dynamics before and after burning in a seminatural grassland of Miscanthus sinensis in Kumamoto, Japan


J. Ryan Stewart, tel. +1 217 265 5461, fax +1 217 244 3469, e-mail: rstewart@illinois.edu


Although fire has been used for several thousand years to maintain Miscanthus sinensis grasslands in Japan, there is little information about the nutrient dynamics in these ecosystems immediately after burning. We investigated the loss of aboveground biomass; carbon (C) and nitrogen (N) dynamics; surface soil C change before and after burning; and carbon dioxide (CO2), methane (CH4), and nitrous oxide (N2O) fluxes 2 h after burning in a M. sinensis grassland in Kumamoto, Japan. We calculated average C and N accumulation rates within the soil profile over the past 7300 years, which were 58.0 kg C ha−1 yr−1 and 2.60 kg N ha−1 yr−1, respectively. After burning, 98% of aboveground biomass and litter were consumed. Carbon remaining on the field, however, was 102 kg C ha−1. We found at least 43% of C was possibly lost due to decomposition. However, remaining C, which contained ash and charcoal, appeared to contribute to C accumulation in soil. There was no difference in the amount of 0–5 cm surface soil C before and after burning. The amount of remaining litter on the soil surface indicated burning appeared not to have caused a reduction in soil C nor did it negatively impact the sub-surface vegetative crown of M. sinensis. Also, nearly 50 kg N ha−1 of total aboveground biomass and litter N was lost due to burning. Compared with before the burning event, postburning CO2 and CH4 fluxes from soil appeared not to be directly affected by burning. However, it appears the short time span of measurements of N2O flux after burning sufficiently characterized the pattern of increasing N2O fluxes immediately after burning. These findings indicate burning did not cause significant reductions in soil C nor did it result in elevated CO2 and CH4 emissions from the soil relative to before the burning event.






carbon dioxide




nitrous oxide


green house gas


water filled pore space


Owing to the steadily rising global demand for energy and the possible exhaustion of petroleum in the coming decades (Owen et al., in press), production of bioenergy from plant biomass has increasingly been considered as a viable alternative source of fuel (Milliken et al., 2007; Heaton et al., 2008). Using bioenergy instead of fossil fuels also could potentially mitigate human-induced global climate change because carbon (C) in bioenergy mainly comes from atmospheric carbon dioxide (CO2). Although maize (Zea mays) grain is currently a major source of biofuel, several plant functional types have been evaluated as candidate bioenergy crops (Heaton et al., 2008). Perennial grassland plant species, which are generally not competitive with food crop production such as switchgrass (Panicum virgatum) or Miscanthus×giganteus, appear to produce more energy effectively while mitigating greenhouse gas emissions (Davis et al., 2009). Grasslands are one of the most abundant land-cover types in the world; they comprise about 41% of global land area, except for areas of permanent ice cover (Adams et al., 1990; White et al., 2000). Grassland ecosystems have high C storage in soil (20–130 Mg C ha−1) and are considered a potential sink or source of C (Carpenter-Boggs et al., 2003; Arshad et al., 2004; Bronson et al., 2004).

Over the past 20–30 years, much interest has been directed toward a highly productive sterile, perennial hybrid grass, M. ×giganteus, for its use as a source of bioenergy (Jones & Walsh, 2001). Miscanthus×giganteus was first collected in Japan in 1935 and then cultivated throughout Europe (Jones & Walsh, 2001). Although natural and seminatural grasslands in Japan, which historically comprised 10% of the land area in Japan in the early 1900s (Imura & Shi, 2004), but in recent years only constitute 4% of the country (Himiyama et al., 1995), are comprised of several graminoid and forb species, including Miscanthus sinensis, which, in addition to Miscanthus sacchariflorus, is one of the parent species of M. ×giganteus. M. sinensis dominates most of these highly diverse grasslands. M. sinensis grasslands, which, in many locations, have been managed for hundreds of years in Japan (Matsumura & Iwata, 1976; Otaki, 1999), comprise about 24% of the grasslands in Japan (National Parks Association of Japan, 1996).

The Miscanthus genus, which utilizes the energy-efficient C4 photosynthetic pathway, (Naidu et al., 2003), is comprised of several species that are considered potential bioenergy crops because of their low-nutrient requirements (Lewandowski et al., 2003; Heaton et al., 2004), high water-use efficiencies (Clifton-Brown et al., 2002), and high productivity (Clifton-Brown et al., 2001; Stewart et al., 2009).

Extensive research in Japan over the past several decades focused on the use of M. sinensis, which is native to Japan, as thatching material for roofs of traditional houses and buildings, organic fertilizer, and livestock feed (Stewart et al., 2009). Also, most of these studies were focused on biomass production and vegetation characteristics in seminatural M. sinensis grasslands. Although this information is useful for better understanding the ecology and agronomy of M. sinensis and even other Miscanthus taxa such as M.×giganteus, there are other factors to consider if information related to this species and its role in the unique seminatural grasslands in Japan can be utilized to further the field of bioenergy both as a source of plant improvement or as a stand-alone bioenergy crop. Also, while considerable effort has been expended in Europe and United States to understand the agronomics of M.×giganteus for the past 20–30 years, most field-based studies have been 10 years or less in duration (Jones & Walsh, 2001). The longest duration of experiments have been, at most, 15 years (Clifton-Brown et al., 2007; Christian et al., 2008).

The unique seminatural M. sinensis grasslands present opportunities to better understand not only the agronomic potential of the Miscanthus genus for long-term production for biomass or biofuel, but also its biogeochemistry. It is estimated burning of M. sinensis grasslands, particularly in the Aso region of Kumamoto, have occurred for more than 10 000 years (Ogura et al., 2002; Miyabuchi & Sugiyama, 2008). Burning as both a natural occurrence and as a management tool in several types of grassland throughout the world has been extensively documented (Zepp et al., 1996; Anderson & Poth, 1998; Towne & Kemp, 2003; Castaldi & Fierro, 2005). Beneficial effects of burning have been demonstrated in grasslands. Zhang et al. (2008) reported increases in light-saturated photosynthetic rate of Agropyron cristatum and Cleistogenes squarrosa in early July in burned grasslands in China compared with those that were not burned. Stem density of grass species also appears to increase due to burning relative to those not burned (Yamamoto et al., 2002; Wang et al., 2006). Knapp (1985) and Blair (1997) found that C4 grasses, such as those found in North American tallgrass prairies, exhibited higher aboveground productivity when burned annually compared with grasses that were not burned.

In M. sinensis grasslands, burning is also important to maintain the vegetational composition of the fire-dependent ecosystem, which is species rich, and to reduce the litter layer to allow for nutrient cycling (Iizumi, 1976; Yamamoto et al., 2002). Yamamoto et al. (2002) also reported that aboveground biomass of M. sinensis in the Aso region of Kumamoto was lower than in nearby grasslands where burning no longer occurs. They also found that where controlled burns no longer occurred, litter was calculated to be three to eight times greater than that found in controlled-burn sites.

To our knowledge, however, no work has been done to characterize C and nutrient dynamics or green house gas (GHG) emissions in M. sinensis grasslands subjected to annual burning events. Because the conditions of grassland ecosystems change drastically after burning, direct and indirect impacts of burning on nutrient cycling in M. sinensis grasslands could be considerably large. Direct impacts of burning, which occur during and just after burning, include C release to the atmosphere by burning of plant biomass and by heating of the surface soil. Indirect impacts of burning, which occur after burning and are often long term in nature (i.e., several months or more), include nutrient loss from the soil surface by weathering and erosion (Fynn et al., 2003). Moreover, increased microbial activity in soil also occurs due to rising soil temperatures and a higher number of wetting and drying cycles (Mills & Fey, 2004; Knicker, 2007). Thus, assessing the direct and indirect impacts of burning to C, N, and GHGs are important for evaluating C and nutrient cycling in M. sinensis grasslands.

In this study, we characterized the direct impacts of burning to the loss of plant biomass; dynamics of C and N; change in mass of C in surface soil; and change of CO2, CH4, and N2O fluxes from the soil surface after burning in a M. sinensis grassland ecosystem, in which annual burning has occurred for more than 40 years, in Kumamoto, Japan. If high amounts of plant material (e.g. ash) remain on the soil surface after burning, this process could contribute to C accumulation in the soil due to the recalcitrant nature of this form of C, which is primarily in the form of charcoal (Lehmann et al., 2006; Marris, 2006). In addition, under certain conditions, we postulate that burning may not negatively impact C sequestration nor result in a change of GHG fluxes at significant levels from within the soil after burning. Outcomes of this study will provide important knowledge to the study of C and nutrient cycle and in the long-term management of M. sinensis grassland ecosystems with the intent that such information could be beneficial to subsequent development of not only M. sinensis as a bioenergy crop, but also other members of the genus.

Materials and methods

Site description

The study was conducted in a semi-natural grassland of M. sinensis located in the northern rim of the Mt. Aso caldera in Kumamoto, Japan (33°01.58′N, 131°03.89′E, 794 m asl) on Cumulic Nonallophanic Andosols (Matsuyama & Saigusa, 1994; Cultivated Soil Classification Committee, 1995) or Melanudans (USDA, 1999). A distinct feature of the soil profile, which was useful in calculating the average C accumulation rate, was the 2AB horizon, commonly known in Japan as K-Ah, which is chemically and visually distinct and was deposited approximately 7300 years ago due to a nearby volcanic eruption (Miyabuchi & Watanabe, 1997).

The terrain of the study site, which was approximately 0.4 ha, was relatively flat and less susceptible to erosion compared with surrounding fields and provided more uniform conditions for measuring nutrient cycling. Physical and chemical properties of the soil are listed in Table 1. Soil C and N amounts on the top 30 cm were 20.2 kg C m−2 and 1.04 kg N m−2, respectively. Soil textures, which varied by depth in the soil profile, were clay loam, loam, or sandy loam (Table 1). Bulk densities of each horizon within the profile ranged from 0.30 to 0.33 g cm−3 (Table 1). Bray(2) P and exchangeable K, which were higher in the top horizon, gradually decreased in deeper horizons (Table 2). The site has a 30-year mean annual precipitation of 3250 mm and air temperature of 9.6 °C.

Table 1.   Physical and chemical characteristics in each soil horizon in a seminatural Miscanthus sinensis grassland in Aso, Kumamoto, Japan
HorizonDepth (cm)Bulk density (g cm−3)Soil texturepHCEC (cmolc kg−1)TC (g C kg−1)TN (g N kg−1)C/NP (Bray 2) (mg P kg−1)Exchangeable K (mg K kg−1)
Sand (%)Silt (%)Clay (%) 
  1. CEC, cation exchange capacity; CL, clay loam; C/N, carbon/nitrogen ratio; K, potassium; L, loam; P, phosphorus; SL, silty loam; TC, total carbon; TN, total nitrogen.

Table 2.   The amount of dry biomass, carbon (C), and nitrogen (N) of plant biomass and litter before and after burning in a semi-natural Miscanthus sinensis grassland in Aso, Kumamoto, Japan
Measured parametersBiomass dry weight (kg ha−1)C (kg ha−1)N (kg ha−1)
  • **

    Statistical differences (P<0.01) between sums of aboveground biomass and litter before and after burning (Mann–Whitney U-test).

  • §

    §After burning, litter and ash could not be separated.

Before burning
 M. sinensis10 82370135249345916.811.3
 Other plant15613978.568.60.951.12
 Subtotal16 197**72387566**354952.2**14.8
After burning
 M. sinensis and other species10212248610.30.38
 Litter and ash§1463454.814.11.380.35
 Loss by burning15 95072407463355050.514.8

Plant species with high relative dominance were M. sinensis (47.1%), Arundinella hirta (9.86%), Pleioblastus argenteostriatus (5.45%), Amphicarpaea bracteata (6.02%), Artemisia indica (3.56%), Lespedeza bicolor (2.99%), Pteridium aquilinum subsp. japonicum (2.77%), and Lespedeza cuneata (2.02%). The only management applied for 40 years and possibly longer was an annual burning event each March. Aboveground biomass at this site was not harvested over the 40-year management period. The study was initiated on 21 March 2009 with a 1-h burn event beginning at 10:00 hours. The fire was ignited by a handheld burner from the leeward side of the study site. Burning and smoke emission ended by 12:00 hours on 21 March.

Aboveground biomass and litter measurements before and after burning

Aboveground biomass and litter of M. sinensis and other plants in the study site were manually cut close to the soil surface before and after burning from 16 randomly selected 1-m2 areas with similar amounts of biomass. Eight of the areas were used for measurements before burning on 26 February and the remaining eight for measurements immediately after burning on 21 March. Litter samples after burning also contained ash because they could not be separated. Samples were oven dried at 70 °C for 48 h, then weighed, ground, and analyzed for C and N with an elemental analyzer (Vario EL III, Elemental, Hanau, Germany).

Amount of C in surface-level soil

Undisturbed soil samples (0–5 c m) were collected with stainless steel cores (100 cm3) from eight replications before and after burning. The cores were divided into the 0–1, 1–2, 2–3, and 3–5 cm depth increments. Bulk density and C content of soil in each layer were calculated (Logsdon & Cambardella, 2000). Soil C concentration was analyzed with an elemental analyzer.

Measurement of CO2, CH4, and N2O fluxes

We measured CO2, CH4, and N2O fluxes with a closed-chamber method described by Toma & Hatano (2007) with four replications before and after burning. Stainless-steel bases were installed 30 min before the first measurement. The bases were removed before burning and reinstalled immediately thereafter. Before burning, gas fluxes were measured at 13:00 hours on 25, 11:00 hours and 16:00 hours on 26 February; 10:00 hours and 14:00 hours on 27 February; 02:00 hours on 28 February; and 14:00 hours on 28 February. Fluxes of GHG measured before burning were considered a daytime baseline. Gas fluxes were measured 2, 4, 7, 11, 23, 31, 45 h after burning from 12:00 hours on 21 March to 7:00 hours on 23 March. A 250-mL gas sample was collected into a tedlar bag (500 mL) for CO2 determination at 0 and 6 min from the time the chambers were closed, which was based on the method of Nakano et al. (2004). Carbon dioxide concentrations were measured with a CO2 analyzer (ZFP-9, Fuji Electric Systems, Tokyo, Japan). Vacuum-10-mL vials sealed with butyl rubber stoppers (SVF-10, Nichiden-Rika, Kobe, Japan) were used to collect 20-mL CH4 and N2O gas samples at 0, 15, and 30 min after chamber closure. Methane and N2O concentrations were determined by a gas chromatograph equipped with a flame ionization detector (GC-8A, Shimadzu, Kyoto, Japan) for CH4 and electron capture detector (GC-14B, Shimadzu) for N2O.

Fluxes of CO2, CH4, and N2O were calculated by the following equation:


where F is the flux (mg m−2 h−1); ρ is the gas density (1.977 × 106 mg m−3 for CO2, 0.717 × 106 mg m−3 for CH4, and 1.978 × 106 mg m−3 for N2O); V is the volume of the chamber (m3); A is the cross-sectional area of the chamber (m2); Δct is the ratio of change in the gas concentration (c) inside the chamber per unit time (t) during the sampling period (m3 m−3 h−1); T is the air temperature ( °C), and α is a conversion factor for CO2 to C (=12/44), CH4 to C (=12/16), or N2O to N (=28/44).

Ancillary measurements

Soil temperature was measured with a thermometer (CT-413WR, CUSTOM, Tokyo, Japan) at 5 cm of soil depth and volumetric soil water content averaged over the top 6 cm of the soil was measured with a moisture sensor (DIK-311D, Daiki Rika Kogyo Co. Ltd., Saitama, Japan) during gas sample collection. Continuous hourly soil temperature measurements were also collected at 5 cm of soil depth from February 26 until the end of the study (Ondotori Jr.TR-52S, T&D Corporation, Nagano, Japan). Total soil porosity and volumetric water content measurements were used to calculate water-filled pore space (WFPS). Air temperature and precipitation information was collected from a nearby weather station.

Statistical analysis

Aboveground biomass, litter, soil C comparisons, and CO2, CH4, and N2O fluxes before and after burning were performed using Student's t-test for normally distributed data and Mann–Whitney's U-test for data not normally distributed as determined by a Chi-square test (Ichihara, 1990).


We calculated the average C and N accumulation rates within the soil profile at the study site over the past 7300 years that was possible given the 2AB horizon. Average C and N accumulation were 58.0 kg C ha−1 yr−1 and 2.60 kg N ha−1 yr−1, respectively.

After burning, aboveground biomass and litter and ash were significantly lower than those before burning (P<0.01) (Table 2). Loss of aboveground biomass and litter due to burning was 98.5% of that before burning. In addition, before burning, C and N in aboveground biomass and litter were significantly higher than aboveground biomass and litter and ash after burning (Table 2). Released C and N in aboveground biomass and litter were 98.6% and 96.7%, respectively, of C and N of aboveground biomass and litter before burning. Only 1.4% and 3.3% of C and N, respectively, of aboveground biomass and litter before burning remained after the burn event (Table 2). Mean C accumulation constituted nearly 57% of remaining C after burning (Table 2). Mass of C before and after burning did not significantly differ in surface soil in the 0–1, 1–2, 2–3, and 3–5 cm layers (Fig. 1).

Figure 1.

 Mass of soil carbon (C) in surface soil (0–1, 1–2, 2–3, and 3–5 cm) before and after burning. Solid squares represent C mass before burning. Open squares represent C mass after burning. Error bars are standard deviations of the means. ns indicates nonsignificant at the P<0.05 level.

When CO2, CH4, and N2O fluxes were measured in February, average of soil temperature at 5-cm depth and WFPS were 8.6 °C and 14.1%, respectively (Fig. 2). Average fluxes of CO2, CH4, and N2O before burning were 72.4 mg C m−2 h−1, −39.0 μg C m−2 h−1, and −0.70 μg N m−2 h−1, respectively (Fig. 2).

Figure 2.

 Changes in air temperature (a), precipitation (a), soil temperature (b), water-filled pore space (WFPS) (c), CO2 flux (d), CH4 flux (e), and N2O flux (f). Error bars are standard deviations of the means. The vertical line indicates when the controlled burn was initiated. The vertical dotted line shows when precipitation began.

As burning commenced from 8:00 hours to 0:00 hours on 21 March, air and soil temperature rose up from 8.4 to 15 °C and 6.08 to 11.7 °C, respectively. Precipitation began from 0:00 hours on 22 March and continued for 18 h. Total accumulated rainfall was 65.5 mm. Over the entire measurement period, CO2, CH4, and N2O fluxes after burning ranged from 35.8 to 111 mg C m−2 h−1, −56.2 to −7.47 μg C m−2 h−1, and −2.08 to 31.4 μg N m−2 h−1, respectively (Fig. 2). However, CO2, CH4, and N2O fluxes beginning at the end of the burn event until precipitation began ranged from 35.8 to 95.3 mg C m−2 h−1, −56.2 to −7.47 μg C m−2 h−1, and 2.59 to 31.4 μg N m−2 h−1, respectively (Fig. 2). Carbon dioxide and CH4 fluxes before and after burning were within similar ranges (Fig. 2). However, average of N2O fluxes 1 day after burning significantly increased from −0.70 to 48.2 μg N m−2 h−1 (P<0.05) (Fig. 2).


Loss of aboveground biomass and biomass C and N by burning

Possibly due to the large spatial variation of vegetative crown size of M. sinensis clumps (Yano and Kayama, 1975), standard deviations of mean values in Table 2 were higher than might be expected. Although most of the aboveground biomass and litter were completely burned, some of the unburned litter remained on the ground after burning (Table 2). Iwanami (1972) reported that 92%–96% of aboveground biomass of M. sinensis and other plant species were consumed by burning in a similar M. sinensis grassland in Kawatabi, Japan.

In a laboratory experiment on effects of burning on biomass, remaining mass of straw of spring wheat (Triticum aestivum), oat (Avena sativa), and flax (Linum usitatissimum) were only 13%, 8%, and 4%, respectively (Heard et al., 2006). Iwanami (1973) reported that when 10 Mg dry wt. ha−1, which is similar in scale to the 10.8 Mg dry wt. ha−1 of M. sinensis measured in our study, was burned in a M. sinensis grassland in Kawatabi, only the aboveground biomass was completely consumed by burning. The belowground biomass only partially burned to a depth of 2 mm. Also, burning occurred only in the outer margins of the vegetative crown. This would indicate the crown of M. sinensis is likely not damaged by burning, which might be advantageous for maintaining the vegetational composition of the M. sinensis in seminatural M. sinensis grasslands (Yamamoto et al., 2002). Kitchen et al. (2009) reported that over a 13-year period, root biomass was greater in annually burned prairies than in those that were not burned in Kansas, United States, which confirms similar work in M. sinensis grasslands in Japan (Iwanami, 1973).

Given that little, if any, belowground biomass C appears to be lost during burning of M. sinensis grasslands, the calculation of C budgets of these ecosystems is possibly simplified. However, detailed studies to measure changes in belowground biomass during burning have not been conducted to completely validate this approach.

Similar to the 98.6% loss of C in aboveground biomass and litter in our study (Table 2), Heard et al. (2006) also reported that loss of C in burned straw of winter wheat, oat, and flax were estimated to be 90.6%, 96.3%, and 96.9%, respectively. Only 102 kg C ha−1 of aboveground biomass and litter was left on the M. sinensis grassland surface (Table 2). Average C accumulation during the 7300-year period, which followed the volcanic eruption, at the rate of 58.0 kg C ha−1 yr−1 was lower than accumulated soil C under 9- (778 kg C ha−1) and 16- (1125 kg C ha−1) year-old fields of M.×giganteus grown on Typic Haplumbrept (Hansen et al., 2004), but higher than that under 2–5-year-old cultivated switchgrass (0 kg ha−1 yr−1), where C inputs equaled outputs (Sartori et al., 2006). Also, Yazaki et al. (2004) reported that in mowed, but unharvested M. sinensis grasslands on Andisols located near Nagano, Japan, the C accumulation rate (3570–4030 kg C ha−1 yr−1) was 62–69 times higher than our calculation. In this study, average C accumulation was 56.9% of remaining C after burning. Therefore, at least 43.1% of C, which remained on the field after burning, was possibly lost to the atmosphere.

Carbon in the form of charcoal, which remained on the soil surface after burning, is known to decompose at slow rates compared with fresh organic C (Seiler & Crutzen, 1980; Skjemstad et al., 2002; Lehmann et al., 2006). Compared with aboveground biomass in M. sinensis grasslands, large amounts of underground biomass are generally stored belowground (Stewart et al., 2009). Carbon in belowground biomass may contribute to C accumulation in soil. However, there is a possibility that charcoal in the residual ash, which was left on the field after burning, may ultimately contribute to C accumulation in soil because of its recalcitrant nature (Lehmann et al., 2006; Marris, 2006).

Heard et al. (2006) also reported that loss of N in burned straw of winter wheat, oat, and flax was estimated to be 98.2%, 99.0%, and 99.8%, respectively. These values are quite similar with what was found in our study (96.7%). Additionally, because mineral N generally increases in soil after burning (Murphy et al., 2006; Knicker, 2007), it is possible that additional N may have been lost through leaching after burning. On the other hand, N accumulation rate was 2.60 kg N ha−1 yr−1. Therefore, addition of N due to biological fixation of N2 by other species, such as L. bicolor and L. cuneata, which are common in M. sinensis grasslands, and wet and dry N deposition possibly contributed to N accumulation in the soil over time. This requires further investigation, but it should be considered that in grassland ecosystems, nearly 20% of the biomass N may come from biological N2 fixation (Pakrou & Dillon, 2000; Kimura et al., 2007). Also, it has been reported that N2-fixing clostridia and nondiazotrophic bacteria symbiotically provide N to M. sinensis plants (Ye et al., 2005; Stewart et al., 2009). Moreover, nearly 10 kg N ha−1 yr−1 of N deposition were reported in agricultural fields in Japan (Hayashi et al., 2007; Kimura et al., 2007).

Carbon in surface soil

Mass of C in the surface soil (i.e., 0–1, 1–2, 2–3, and 3–5 cm depths) did not significantly change after burning (Fig. 1). Since some litter remained unburned, it is unlikely the burning temperature and/or short duration of burning were sufficient to have an impact on reducing soil C. Increases in soil temperature in surface soil at burning are generally minimal in this grassland ecosystem even if temperatures at the air–surface interface are >100 °C (Iwanami, 1972). Iwanami (1972) reported that soil at 1 cm depth increased in the range of 8–14 °C when compared with the temperature before burning when 3–5 Mg dry wt. ha−1 was burned in a M. sinensis grassland in Kawatabi. In this study, soil temperature at the 5-cm depth increased from 7.54 to 10.6 °C during the 1-h burn period (Fig. 2). This increase followed the same pattern as ambient air temperature and likely indicates temperature changes were due to increased sunlight radiation, but not to burning. Although CO2 fluxes from the soil were not measured during burning, if it is assumed there was a lag time in the amount of organic matter decomposition and concomitant increases in CO2 fluxes due to increased temperature, the data collected 2 h after burning indicate very small fluxes occurred. At 12:00 hours, CO2 flux was 95.3 mg C m−2 h−1. This corresponds to about 0.01% of soil C in the top 5 cm of soil and indicates that burning did not directly disturb soil C in the surface soil. As mentioned, burning does not appear to be detrimental to C accumulation in surface soils in this grassland system.

Change of CO2, CH4, and N2O fluxes after burning

The CO2 flux, which included microbial and root respiration, during the winter was lower than the values observed in a M. sinensis grassland growing on Andisols near Nagano in November (200–300 mg C m−2 h−1). The CH4 flux was nearly within the same range (approximately −30 to 0 μg C m−2 h−1) as that in an unfertilized and unburned orchardgrass pasture during winter on a volcanic ash soil in Nasu, Japan (Mori et al., 2005). Nitrous oxide was a little lower than the value (0–25 μg N m−2 h−1) of Jorgensen et al. (1997) in a M. ×giganteus field on Typic Haplumbrept in Denmark.

When measurement of CO2, CH4, and N2O fluxes began at 12:00 hours, burning and smoke emission had ended. Because of a rain event that began at 0:00 hours on 22 March, only flux data collected before that time was analyzed given rain can drastically change soil conditions due to changes in temperature and moisture. Considering several factors influence GHG flux from soil (Conrad, 1995; Bouwman et al., 2001), caution is needed when comparing measured GHG fluxes of other grassland ecosystems where burning occurs. After burning, there were no significant differences in CO2 and CH4 fluxes before and after burning. However, N2O fluxes 1 day after burning appear to have been affected by burning. Only N2O fluxes significantly increased compared with that before burning (P<0.05).

Besides that reported in our study, there appears to be little, if any, information about changes in GHG fluxes from soil before and after burning in a grassland ecosystem. Levine et al. (1996) reported no differences in N2O fluxes before and after burning in savannas in South Africa. However, our results may indicate frequent measurement of N2O fluxes at least 1 day after burning might be needed to more accurately quantify N2O emission after burning, particularly if the increase of N2O fluxes after burning is higher than that of mean annual N2O fluxes. The range of N2O fluxes after burning (2.59–31.4 μg N m−2 h−1) was similar to values measured over the period of a year in an unfertilized and unburned orchardgrass pasture on a volcanic ash soil in Nasu (1–10 μg N m−2 h−1) (Mori et al., 2005) and that measured from April to July in a M. ×giganteus field on Typic Haplumbrept in Denmark (0–50 μg N m−2 h−1) (Jorgensen et al., 1997). Although future investigations of longer duration will be required to determine a more accurate estimate of the contribution of gas emissions due to burning, burning may not directly cause substantial increases in CO2, CH4, and N2O fluxes as they relate to total annual gas emissions.


We would like to thank Mr. Makoto Nakabo and his staff of the Kyushu Biomass Forum for providing technical assistance in our field research in Aso, Kumamoto, Japan. We also would like to thank Mr. Syotaro Kuwabara of Gifu University for providing vegetational analysis at the research site. This research was supported by the Energy Biosciences Institute.