Seasonal variation and fire effects on CH4, N2O and CO2 exchange in savanna soils of northern Australia

Agricultural and Forest Meteorology 151 (2011) 1440–1452

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Seasonal variation and fire effects on CH4 , N2 O and CO2 exchange in savanna soils of northern Australia Stephen J. Livesley a,b,∗ , Samantha Grover c , Lindsay B. Hutley c , Hizbullah Jamali b , Klaus Butterbach-Bahl d , Benedikt Fest b , Jason Beringer c , Stefan K. Arndt b a

School of Geography and Environmental Science, Monash University, VIC, Australia Department of Forest and Ecosystem Science, The University of Melbourne, VIC, Australia c School of Environmental and Life Sciences, Charles Darwin University, NT, Australia d Institute for Meteorology and Climate Research, Karlsruhe Institute of Technology, Kreuzeckbahnstr 19, 82467 Garmisch-Partenkirchen, Germany b

a r t i c l e

i n f o

Article history: Received 14 August 2010 Received in revised form 29 December 2010 Accepted 2 February 2011 Keywords: Methane Nitrous oxide Carbon dioxide Nitrification Fire Savanna woodland Termites Soil gas diffusion

a b s t r a c t Tropical savanna ecosystems are a major contributor to global CO2 , CH4 and N2 O greenhouse gas exchange. Savanna fire events represent large, discrete C emissions but the importance of ongoing soilatmosphere gas exchange is less well understood. Seasonal rainfall and fire events are likely to impact upon savanna soil microbial processes involved in N2 O and CH4 exchange. We measured soil CO2 , CH4 and N2 O fluxes in savanna woodland (Eucalyptus tetrodonta/Eucalyptus miniata trees above sorghum grass) at Howard Springs, Australia over a 16 month period from October 2007 to January 2009 using manual chambers and a field-based gas chromatograph connected to automated chambers. The effect of fire on soil gas exchange was investigated through two controlled burns and protected unburnt areas. Fire is a frequent natural and management action in these savanna (every 1–2 years). There was no seasonal change and no fire effect upon soil N2 O exchange. Soil N2 O fluxes were very low, generally between −1.0 and 1.0 ␮g N m−2 h−1 , and often below the minimum detection limit. There was an increase in soil NH4 + in the months after the 2008 fire event, but no change in soil NO3 − . There was considerable nitrification in the early wet season but minimal nitrification at all other times. Savanna soil was generally a net CH4 sink that equated to between −2.0 and −1.6 kg CH4 ha−1 y−1 with no clear seasonal pattern in response to changing soil moisture conditions. Irrigation in the dry season significantly reduced soil gas diffusion and as a consequence soil CH4 uptake. There were short periods of soil CH4 emission, up to 20 ␮g C m−2 h−1 , likely to have been caused by termite activity in, or beneath, automated chambers. Soil CO2 fluxes showed a strong bimodal seasonal pattern, increasing fivefold from the dry into the wet season. Soil moisture showed a weak relationship with soil CH4 fluxes, but a much stronger relationship with soil CO2 fluxes, explaining up to 70% of the variation in unburnt treatments. Australian savanna soils are a small N2 O source, and possibly even a sink. Annual soil CH4 flux measurements suggest that the 1.9 million km2 of Australian savanna soils may provide a C sink of between −7.7 and −9.4 Tg CO2 -e per year. This sink estimate would offset potentially 10% of Australian transport related CO2 -e emissions. This CH4 sink estimate does not include concurrent CH4 emissions from termite mounds or ephemeral wetlands in Australian savannas. © 2011 Elsevier B.V. All rights reserved.

1. Introduction The increased concentration in the atmosphere of greenhouse gases including carbon dioxide (CO2 ), methane (CH4 ) and nitrous oxide (N2 O) as a result of anthropogenic activity is widely recognised as the cause of global climate warming and an increase in the

∗ Corresponding author at: Dept. of Resource Management and Geography, The University of Melbourne, Burnley campus, Richmond, Melbourne, VIC 3121, Australia. Tel.: +61 439 615 772. E-mail address: [email protected] (S.J. Livesley). 0168-1923/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.agrformet.2011.02.001

intensity and frequency of extreme weather events (IPCC, 2007). These gases are all naturally present in the atmosphere and cycle between the land, the ocean and the atmosphere over a range of time scales. Global climate models have been developed as tools to predict global biogeochemical cycling and climate systems, but their efficacy is limited by our understanding of the underlying processes which drive the cycling of greenhouse gases (Beringer et al., 2011), this is particularly so for CH4 and N2 O. Savanna ecosystems are a significant global biome and cover approximately one sixth of the earth’s land surface and have been estimated to produce 30% of the earth’s total net primary productivity (Grace et al., 2006). Tropical savanna ecosystems have been

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identified as a major source of trace gas emissions, however seasonal variation and uncertainty in the magnitude of these emissions is considerable (Bousquet et al., 2006; Crutzen and Andreae, 1990; Grace et al., 2006; Potter et al., 1996). The exchange of CO2 , CH4 and N2 O in natural savanna ecosystems is predominantly a function of: • • • • • • •

gross primary productivity (CO2 ) plant and soil respiration (CO2 ) termite activity (CH4 ) periods of wetland inundation and water level (CH4 , N2 O) soil methane oxidation (CH4 ) soil nitrification and denitrification (N2 O) fire events (CO2 , CH4 and N2 O)

Savannas are characterised by a grass understorey with a tree/shrub overstorey in regions with distinct wet and dry seasons and frequent low-intensity fires (every 1–5 years) through natural or human induced actions (Grace et al., 2006; Hutley and Beringer, 2010). Savanna fires are integral to nutrient cycling and the life cycle of many plants. Global climate models generally perform poorly at simulating savanna ecosystems (Hutley et al., 2005), probably as they have been developed from process understanding of principally temperate ecosystems (Archibald et al., 2009). In temperate ecosystems, plant productivity, or plant C exchange, is primarily limited by temperature and light availability, whereas in savanna ecosystems water availability and nutrient cycling primarily limit productivity (Archibald et al., 2009). The high spatial variability and complex interaction of biotic–abiotic factors that influence plant productivity in Australian savannas are a major challenge to our ability to simulate savanna-atmosphere exchange (Kanniah et al., 2011). Global climate models have recently incorporated fire as a disturbance factor (Thonicke et al., 2010), but principally to simulate rare, high-intensity fires typical of temperate ecosystems (Spessa et al., 2005). Savanna fires are a major global source of aerosols, trace gases and volatile organic compounds and direct research studies have been carried out in Africa (Cofer et al., 1996; Grace et al., 2006; Otter and Scholes, 2000; Sinha et al., 2004), South America (Crutzen and Andreae, 1990) and in Australia (Beringer et al., 2003; Hurst et al., 1994; Meyer et al., 2008; Russell-Smith et al., 2009; Shirai et al., 2003). It has been estimated that tropical ecosystems represent a 12–14 Tg N2 O source (Crutzen and Andreae, 1990) but only 0.1–0.3 Tg being produced through pyro-denitrification (loss of N due to fire). Nitrous oxide can be produced in aerobic soils as a by-product of the microbial nitrification process whereby NH4 + is transformed to NO3 − , or as an integral process step in anaerobic denitrification processes, where NO3 − , NO2 − or NO is transformed to N2 O and then N2 (Conrad, 1996; Firestone and Davidson, 1989). The strong seasonality of savanna rainfall and soil moisture can create significant seasonal pulses of soil N2 O (Davidson et al., 1993) such that tropical savannas may contribute up to 4.4 Tg N2 O y−1 (Castaldi et al., 2006). However, the importance of soil-atmosphere N2 O exchange in savanna ecosystems is poorly understood because of the paucity of data and limited process understanding in N-limited savanna soils (Rees et al., 2006). Oxidation of CH4 by soil methanotrophic bacteria is the largest CH4 sink in the terrestrial biosphere (Dalal et al., 2008; King, 1997; Potter et al., 1996). It has been estimated that 17–40 Tg of CH4 is removed from the atmosphere through soil methane uptake (oxidation) each year (Potter et al., 1996; Reeburgh et al., 1993). Approximately 40% of this is taken up by soils of warm-dry ecosystems such as savanna, semi-arid steppe, seasonally dry tropical forest and chaparral (Potter et al., 1996) of which savannas are by far the largest of these warm-dry ecosystems. There are large uncertainties and inter-annual variability associated with estimates of the global CH4 balance, principally with regards to ephemeral wet-

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lands and wildfire emissions (Bousquet et al., 2006) but also the paucity of data on soil CH4 uptake (oxidation) in warm-dry ecosystems (Otter and Scholes, 2000; Potter et al., 1996). Savanna soils may also act as a CH4 source through subterranean termite activity and soil methanogenesis under anoxic conditions, such that soil CH4 oxidation may be offset or there is net soil CH4 emission (Otter and Scholes, 2000; Poth et al., 1995; Zimmerman et al., 1982). Other savanna sources CH4 are termite mounds and tree-dwelling termite colonies (Jamali et al., 2011) as well as swamps and ephemeral wetlands (Otter and Scholes, 2000). Fire may affect soil-atmosphere exchange of greenhouse gases after the combustion event (Castaldi et al., 2010, 2006) by altering soil C inputs, nutrient inputs, surface microbial activity, surface moisture and temperature. However, the effect of fire upon savanna soil-atmosphere exchange of N2 O, CH4 and other trace gases is unclear and contradictory (Anderson and Poth, 1998; Castaldi et al., 2006; Pinto et al., 2002). Savanna ecosystems cover about one quarter of the Australian continent. A great deal of progress has been made towards ecological, eco-physiological and C cycle understanding in savanna woodlands of the Northern Territory, principally at a single flux tower at Howard Springs (Beringer et al., 2003; Chen et al., 2003; Hutley et al., 2005; O’Grady et al., 1999). However, there have been no studies of soil-atmosphere exchange of N2 O or CH4 in an Australian savanna ecosystem. The objectives of this study were to (i) to study the seasonal variability of soil-atmosphere exchange of CO2 , CH4 and N2 O over an entire wet–dry seasonal cycle in a mesic savanna woodland in north Australian where ecosystem scale CO2 fluxes have been, and are being, measured continuously through eddy covariance; (ii) to quantify the effects of fire upon soil greenhouse gas exchange in savanna soils and (iii) to investigate the relationship between soil CO2 , CH4 and N2 O flux and dynamic soil physiochemical properties to understand the driving environmental factors and processes so as to aid the future development of process-based models of savanna biogeochemical function.

2. Methods 2.1. Study site This study was located at the Howard Springs eddy covariance site, 35 km south east of Darwin, NT (Fig. 1B). The climate is wet–dry tropical, with a highly seasonal rainfall and relatively aseasonal air temperature. The Howard Springs weather station (Australian Bureau of Meteorology since 1982) located ∼20 km from the flux tower has a mean rainfall of 1782 mm per annum, compared with the mean rainfall at the Howard Springs tower site of 1824 mm (2001–2006). Mean monthly maximum air temperature of the region varies by only 2 ◦ C over the year, although minimum temperatures show more seasonality. May to September (incl.) are typically dry season months of no rainfall and decreasing soil moisture content. October and November are typical ‘dry–wet transition’ months, when air humidity and temperature increase, annuals germinate in response to sporadic, small–medium rainfall events. December to April (incl.) are typically wet season months. Howard Springs is a mosaic of Eucalypt-dominated woodland, open-forest savanna, closed forest and seasonally flooded swamps and wetlands. Open-forest savanna is the main vegetation type (∼80% area) dominated by an over-storey (14–16 m high) of Eucalyptus tetrodonta (F. Muell.) and Eucalyptus miniata (Cunn. Ex Schauer). Over-storey tree stem density was low at 156 ± 28 (SE). The understorey consists of semi-deciduous and deciduous small trees and shrubs, but is dominated by tall C4 grasses, such as the annual species Sorghum spp. and the perennial grass Heteropogon triticeus. All flux measurements were made in open-forest savanna.

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Fig. 1. Location of the sampling site at the Howard Springs flux tower, NT (−12.494341 latitude, 131.152298 longitude) in the seasonal wet/dry tropics of northern Australia.

Soils are red-earth Kandosols (Isbell, 2002). Open savanna forest dominated by E. tetrodonta and E. miniata is the most widespread tall-grass savanna types in north Australia, representative of the frequently burnt savanna that extend from north-western Western Australia to the Gulf of Carpentaria in northern Queensland. A full description of the savanna floristics of the region is given by Hutley et al., 2011. 2.2. Continuous soil gas exchange measurements Greenhouse gas fluxes were measured continuously for fifteen months from October 2007 to December 2009 using an automated trace gas measurement system (Fest et al., 2009), consisting of a gas chromatograph (GC, SRITM , Torrance, CA, USA) linked to ten acrylic chambers with opaque white lids that open and close automatically through pneumatic piston controls. Two chambers had an internal headspace volume of 24.0 L, four chambers were 30.4 L and four chambers were 37.5 L. Two base frames for each chamber were inserted ∼0.05 m into the soil and chambers attached using clamps and closed cell foam. Chambers were moved from one frame to the other on a weekly basis to prevent chamber-feedback effects. Two of the ten chambers were attached to 1.0 m tall frames enclosing termite mounds (data not presented). Chamber lids opened automatically when rainfall exceeded 0.6 mm in a 5 min period, as detected by a tipping bucket rain gauge. To measure gas fluxes, chambers 1–5 would close for 2 h (or 1 h 20 min after September 2008), then chambers 1–5 would open and chambers 6–10 would close. During chamber closure, four air samples (∼0.5 L each) were withdrawn at equal time intervals. Each air sample took 4 min to withdraw at approximately 200 mL min−1 passing through an IRGA for continuous CO2 concentration measurement (LI820, LICOR Inc., Lincoln, USA). The increase in CO2

concentration between air sample 1 and 2 was used to calculate soil respiration. At the end of a 4 min sampling withdrawal, two 3 mL loops collected and passed chamber air into the GC for measurement of N2 O concentration using a 63 Ni electron capture detector (ECD) and CH4 concentration using a flame ionisation detector (FID). The GC was calibrated with certified CO2 , N2 O and CH4 concentration standard gas (Air Liquide, USA) every 30 min. Nitrous oxide and CH4 fluxes were calculated from a linear regression of concentration change. Six flux measurements were made for each chamber in a 24 h period. Soil moisture and soil temperature at a depth of 0.10 m was continually measured by Theta probe (DeltaT, UK) at the eddy covariance tower within 500 m of the trace gas measurement system. Rainfall was recorded using a tipping bucket rain gauge. The minimum detectable limits (MDL) were calculated as: MDL = 2 × SD × V × A × T where SD is the standard deviation of the gas concentration in ambient air, V is the volume of the chamber (L), A is the area of soil under the chamber (m2 ) and T is the time for incubation (h). For the automated chamber and van mounted GC measurements, the MDL for N2 O was 0.77 ␮g N m−2 h−1 and for CH4 was 1.64 ␮g C m−2 h−1 , whereas for the manual chamber and laboratory GC measurements, the MDL for N2 O was 0.05 ␮g N m−2 h−1 and for CH4 was 0.07 ␮g C m−2 h−1 . Howard Springs typically experiences fire annually or biannually. A moderate intensity fire (∼3 MW m−1 ), typical of mesic, tall-grass savanna, was lit near the trace gas measurement system on 27 August 2008 (late dry season). Chambers were removed just before the burn and returned to the same base frames within 2 h of the fire passing. After the fire, three chambers were in a burnt area and four chambers were in an unburnt area protected by a fire

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break. Soil samples (0–0.05 m) were collected from these burnt and unburnt areas at two week intervals and analysed for soil inorganic N concentration as described in Section 2.4. Unplanned small wild fires also occurred at the site on 30 April 2008 and 25 June 2008 but did not affect the flux measurements presented.

analyser (EuroEA 3028HT, EuroVector SpA, Milan, Italy) connected to an isotope ratio mass spectrometer (IsoPrime, GV Instruments, Manchester, UK). Soil volumetric moisture content was estimated according to soil core volume and converted to percent water filled pore space (%WFPS) according to:

2.3. Episodic soil gas exchange measurements %WFPS = The manual closed chamber method (Hutchinson and Mosier, 1981) was used to provide a complementary and spatially representative assessment of GHG fluxes in burnt and unburnt plots. These manual chamber measurements were made in six plots within 1.5 km of the continuous soil trace gas exchange measurements. Three burnt plots were established in an experimental, late dry season, fire on 30 September 2007, whilst three unburnt plots had fire excluded during the 2007 dry season. Manual chambers were made of PVC pipe (diameter 0.25 m, height 0.245 m, volume 12.0 L, basal area 0.049 m2 ) with a twist-lid (PVC) incorporating a butyl-rubber septum and a rubber O-ring to form a gas tight seal. Manual chamber bases were installed at least 1 h before lid closure, and installed at different locations within a plot on each measurement occasion to enable soil sampling from directly beneath after flux measurement. Five chambers were installed at each plot, on each measurement occasion. The litter layer was cut around the circumference of each chamber base and the chambers were inserted approximately 0.02 m into the soil surface. Once the manual chamber lids were attached and twist-sealed, 20 mL headspace gas samples were taken at 0, 20, 40 and 60 min after closure using a 20 mL syringe (TerumoTM USA) and a one-way stopcock. Gas samples were stored in exetainers were usually analysed within 20 days for N2 O and CH4 by gas chromatography (Schimadzu GC17A with N 2 carrier gas). Injection of a single gas sample filled two 2.0 mL sample loops, one leading to a flame ionisation detector for the determination of CH4 concentration and one leading to an electron capture detector for the determination of N2 O concentration. Exetainers were stored in a cool, dark container and over-pressurised (20 mL in 12 mL container volume) to ensure any minor leaks were from the exetainers to the bulk air. Soil N2 O and CH4 fluxes were calculated from the linear increase or decrease in concentration with time. Soil CO2 emission rates for each chamber were measured after headspace gas sample collection was complete, using an Infra-red gas analyser (IRGA, EGM, PP-SystemsTM , UK). The individual manual chamber lids were removed and the chambers vented. A two-port chamber lid was then attached and the linear increase in CO2 concentration instantaneously measured between 90 and 180 s after lid closure using the IRGA in a closed dynamic setup. Soil temperature at 0.03 m depth and soil water content were measured at each chamber at sampling. 2.4. Soil sampling and measurements Surface litter was collected from each manual chamber and oven-dried at 70 ◦ C for 48 h to determine dry mass. Soil was sampled from each chamber on each measurement occasion using soil cores (diameter, 72 mm). Soil samples were transferred immediately to an ice box (∼4 ◦ C) and then stored refrigerated for 1–4 days prior to analysis. The soil samples were weighed, and subsamples were removed for analysis of gravimetric water content and soil NO3 − and NH4 + . Soil samples were extracted with 1 M KCl (1:4, soil:KCl) and shaken for 1 h on a flat-bed shaker, then filtered (Whatman 42) and frozen prior to analysis of NO3 − and NH4 + on a TechniconTM Auto-analyser. Gravimetric water content was determined through oven drying at 105 ◦ C for 48 h. The remaining soil was air-dried, sieved to determine stone content and then archived for C and N measurements. A sub-sample of each air-dried soil sample was analysed for total C and N content using an elemental

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 1 − (BD/PD)

where  is the volumetric soil water content (%), BD is the bulk density (kg m−3 ) and PD is the particle density constant (0.265 kg m−3 ). 2.5. Soil gas diffusivity In the dry season of July 2009 soil gas diffusion was measured at contrasting soil moisture contents according to von Fischer et al. (2009) from the decrease in concentration of sulphur hexa fluoride (SF6 ) after a 12 mL volume had been injected into a manual chamber headspace. Each manual chamber was inserted ∼0.02 m into the soil and was fitted with a small fan to maintain good mixing of SF6 within the headspace. Three contrasting soil moisture contents were established: (i) dry soil (not irrigated), (ii) medium soil moisture (50 L m−2 added one day before) and (iii) wet soil moisture (50 L m−2 added one day before and 20 L m−2 added in the hour before). Three replicate gas flux and diffusion measurements were made in each water treatment (dry, medium, wet) at three replicate locations (three of the unburnt plots used for manual chamber measurements). Twenty milliliter gas samples were taken from the chamber headspace at 2, 12, 22 and 32 min after chamber closure using a 20 mL syringe (Terumo, USA) and a one-way stopcock. No sampling in the initial 2 min enabled equilibration and good mixing of SF6 . Gas samples were immediately transferred, stored in preevacuated exetainers and analysed for CH4 and N2 O concentration as described for episodic manual chamber measurements. 2.6. Soil N mineralisation Soil nitrification and ammonification were estimated by direct incubation of nylon mesh bags containing 4.0 g of ion-exchange resin beads (Dowex Marathon MR-3, Michigan, USA) at a depth of ∼0.05 m. Three replicate resin bags were buried in the three burnt and three unburnt replicate plots on six separate occasions during the 15 month measurement period. The resin bags were collected after a soil incubation period of between 23 and 98 days and then stored refrigerated (∼4 ◦ C) before extraction. Total exchangeable ions were extracted from the resin with four repeat 1 M KCl washes (1:25, resin:KCl) that were frozen prior to analysis for NO3 − and NH4 + concentration on a TechniconTM segmented flow autoanalyser. 2.7. Data analysis Near-continuous soil fluxes were measured using the automated chamber system over a 15 month period, with mean fluxes estimated from eight automated chambers (two measured termite mounds; data not presented) during the first measurement period (October 2007–May 2008). In the second measurement period (July 2008–December 2008) mean fluxes were estimated from three chambers in the burnt area and four chambers in the unburnt area (one of the eight chambers was damaged in a wildfire event in July 2008). Mean soil fluxes from manual chamber measurements were estimated from 15 manual chambers per treatment (burnt and unburnt), five chambers in each of the three replicate plots. CH4 and N2 O data from the manual chamber measurements were normally

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Table 1 The concentration of total carbon (C) and nitrogen (N) and Bray extractable phosphorus (P) from burnt and unburnt soils at Howard Springs. Standard errors from three replicate soil pits are in parentheses. No significant differences between burnt and unburnt treatments (Student t-test). Depth (m) C (%)

0–0.1 0.1–0.2 0.2–0.3 0.4–0.5 0.7–0.8 0.9–1.0

P (mg kg−1 )

N (%)

BD (g cm−3 )

pH (1:5 water)

EC (␮s cm−1 )

Unburnt

Burnt

Unburnt

Burnt

Unburnt

Burnt

Unburnt

Burnt

Unburnt

Burnt

Unburnt

Burnt

2.34 (0.08) 1.32 (0.05) 0.85 (0.03) 0.67 (0.03) 0.36 (0.02) 0.21 (0.02)

2.09 (0.06) 1.04 (0.04) 0.90 (0.03) 0.68 (0.02) 0.38 (0.02) 0.26 (0.02)

0.08 (0.01) 0.05 (0.01) 0.03 (0.00) 0.03 (0.00) 0.02 (0.00) 0.02 (0.00)

0.06 (0.01) 0.04 (0.00) 0.03 (0.00) 0.02 (0.00) 0.02 (0.00) 0.02 (0.00)

2.06 (0.47) 1.16 (0.09) 0.75 (0.06) 0.72 (0.21) 1.33 (0.11) 1.45 (0.22)

2.78 (0.57) 1.30 (0.20) 0.89 (0.17) 1.13 (0.30) 1.49 (0.29) 0.83 (0.28)

1.33 (0.07) 1.39 (0.08) 1.41 (0.06) 1.45 (0.09) 1.52 (0.09) 1.57 (0.08)

1.46 (0.06) 1.46 (0.05) 1.45 (0.01) 1.46 (0.06) 1.54 (0.05) 1.61 (0.04)

5.88 (0.02) 5.50 (0.11) 5.32 (0.05) 5.39 (0.06) 5.43 (0.09) 5.49 (0.02)

5.74 (0.13) 5.38 (0.05) 5.38 (0.06) 4.99 (0.32) 5.50 (0.06) 5.49 (0.08)

18.51 (1.48) 16.98 (0.57) 20.55 (2.25) 16.09 (1.07) 13.52 (0.38) 13.03 (2.01)

18.73 (1.83) 17.51 (0.96) 16.22 (3.24) 16.65 (2.08) 15.99 (3.29) 17.79 (3.37)

distributed, whereas CO2 data were not and required log transformation to normalise distribution prior to statistical analysis. Statistical significance was defined as the difference at the 95% level (i.e. p ≤ 0.05). The effect of fire was investigated using analysis of variance (ANOVA) to compare fluxes from burnt and unburnt sites. The effect of seasonality was investigated using an ANOVA to compare fluxes from each measurement occasion. Tukey’s post hoc test was used to determine which occasions differed from one another. Stepwise multiple linear regression was used to investigate relationships between gas fluxes and the environmental drivers of soil water content, soil temperature, bulk density, nitrate and ammonia from soil, and nitrate and ammonia from resin bags. The r2 indicated the amount of variation in trace gas flux that can be explained by changes in these environmental variables. The effect of water content on diffusivity and fluxes of CH4 and N2 O was investigated using an ANOVA, and Tukey’s post hoc test was applied where ANOVA indicated significant difference. Monthly mean flux rates (mass CO2 /CH4 /N2 O ha−1 d−1 ) were estimated for the automated chamber data of unburnt savanna soil and manual chamber data of burnt and unburnt savanna soil. For those months without chamber measured data, flux was inferred from an average of the mean flux of the preceding and antecedent month (measured or inferred). Monthly mean flux were multiplied by the days in each month and summed to provide an estimate of annual CO2 , CH4 and N2 O flux. Annual flux in carbon dioxide equivalents was estimated for each GHG based on global warming potentials (GWP) of 1 (CO2 ), 25 (CH4 ) and 296 (N2 O) (Solomon et al., 2007).

nificantly less in the burnt plots as compared to the unburnt plots (Table 2), markedly so in the first few months after the fire. There was no apparent effect from the September 2007 fire on topsoil NH4 + and NO3 − concentrations, although the first three measurements after the event were lost as a result of processing errors (Fig. 2). As an indication of spatial variation, the monthly %CV for soil NO3 − was between 23 and 295, and for NH4 + was between 25 and 256. Soil NH4 + accumulated in the 2008 dry season and decreased as the 2008/09 wet season developed, whereas, soil NO3 − concentrations were small (≤1.0 mg N kg−1 ) throughout. After the controlled burn on 27 August 2008, there was significantly greater soil NH4 + concentrations in the burnt area (n = 5) than the unburnt area (n = 5), although by December the NH4 + concentrations were similar (Fig. 3). Potential nitrification measured by ion-exchange resin bag incubation was greatest at the break of rains in October/November 2007, and was greater in plots that had been recently burnt (Fig. 4). There was no apparent nitrification throughout the late wet season and dry season in 2008. In contrast, the potential ammonification rates were between 100 and 400 ng N g−1 resin d−1 throughout the 16 month measurement period, with no apparent difference according to season or fire treatment (Fig. 4). There was significantly less surface organic litter in the burnt plots after the September 2007 controlled burn, however, by the early wet season measurements in December 2008 this difference was less apparent.

3. Results

Soil-atmosphere fluxes of N2 O from manual chambers were very small (−1.0 to 1.0 ␮g N m−2 h−1 ), beyond the detection limit of the GC. There was no significant effect from fire (Fig. 2 and Table 2), such that overall the mean soil N2 O flux was greatest at 0.21 ␮g N m−2 h−1 in the wet season and least in the late dry season and ‘dry–wet transition’ such that the N2 O flux was negative, −0.20 ␮g N m−2 h−1 (Fig. 2 and Table 2). As an indication of spatial variation, the %CV of monthly manual chamber N2 O flux measurements was between 52 and 7209. The near-continuous measurement of soil N2 O flux by automated chambers showed similarly small fluxes. There was little

3.1. Fire effects on soil properties There was no significant difference in total soil C, N and Bray P in 1.0 m profiles between burnt and unburnt plots 21 months after the controlled fire event (September 2007). Total soil C reached up to 2.3% at 0–0.10 m, Bray P reached up to 2.8% at 0–0.10 m, whereas total soil N was <0.1% throughout the 1.0 m profile (Table 1). Following the controlled moderate-intensity fire in September 2007, soil temperature was significantly greater and soil WFPS sig-

3.2. Seasonal pattern of soil gas exchange and the effect of fire

Table 2 Summary ANOVA table of season and fire effects upon soil-atmosphere trace gas exchange and soil environmental variables (upper 50 mm) as measured on eight occasions through manual chamber incubations in burnt and unburnt plots at the Howard Springs savanna woodland, NT, Australia. Positive values indicate emissions and negative values soil uptake.

CO2 (mg C m−2 h−1 ) N2 O (␮g N m−2 h−1 ) CH4 (␮g C m−2 h−1 ) Soil temp (◦ C) WFPS (%) NH4 + (mg kg−1 ) NO3 − (mg −1 )

Season p value

Dry Mean (SE)

Dry–wet transition Mean (SE)

Wet Mean (SE)

Fire p value

Burnt Mean (SE)

Unburnt Mean (SE)

0.000 n.s. n.s. 0.000 0.000 0.000 0.003

131.2 (21.0) a −0.30 (0.17) −12.80 (1.83) 32.7 (0.52) a 5.29 (0.30) a 2.69 (0.29) ab 0.50 (0.09) a

102.1 (6.5) a −0.20 (0.25) −12.73 (1.78) 37.74 (0.45) b 12.83 (0.81) b 18.68 (6.57) a 0.43 (0.08) a

237.7 (13.2) b 0.21 (0.13) −10.37 (1.24) 31.19 (0.23) c 45.50 (1.27) c 2.29 (0.17) b 0.05 (0.03) b

n.s. n.s. n.s. 0.000 0.021 0.022 n.s.

166.8 (14.8) 0.09 (0.17) −12.41 (1.30) 35.0 (0.46) a 22.96 (1.93) a 5.42 (2.38) 0.25 (0.04)

153.7 (10.3) −0.1 (0.15) −11.30 (1.34) 33.1 (0.36) b 26.42 (2.01) b 6.03 (1.52) 0.36 (0.08)

Significant differences (p < 0.05) among seasons, or between burnt and unburnt treatments, indicated by contrasting lower case letters (a, b or c).

S.J. Livesley et al. / Agricultural and Forest Meteorology 151 (2011) 1440–1452

Soil temperature (°C)

45

35 30

Soil WFPS (%) Organic litter (g m-2)

NH4+ (mg N kg-1) NO3- (mg N kg-1)

N2O (µg N m-2 h-1)

CH4 (µg C m-2 h-1)

A

40

25 60 50 40 30 20 10 0 1,500 1,250 1,000 750 500 250 0 20

burnt

unburnt

burnt

unburnt

B

C

D

15 10 5 0

1

E

0.5 0 -0.5 1

F

0.5 0 -0.5 -1 -1.5 0 -5

G

-10 -15 -20 -25

CO2 (mg C m-2 h-1)

1445

300

H

burnt

unburnt

200 100 0 1 Oct 07

20 Dec 07

9 Mar 08

28 May 08

16 Aug 08

4 Nov 08

23 Jan 09

Fig. 2. Soil physiochemical properties (0–0.05 m) and soil CH4 , N2 O and CO2 exchange in three plots burnt in September 2007 and three unburnt plots on eight occasions within a 16 month period between October 2007 and January 2009. Presented values (with standard errors) are the mean of three plots; each plot being the mean of five separate chamber measurements.

variation in N2 O flux according to season or in response to controlled fire event (27 August 2008). The mean soil N2 O flux rate in unburnt savanna was 0.287 ± 0.024 ␮g N m−2 h−1 in the first measurement period (n = 8), and 0.019 ± 0.040 ␮g N m−2 h−1 in the

second measurement period (Fig. 5). The MDL for individual N2 O flux measurements was 0.7 ␮g N m−2 h−1 which exceeded many measured fluxes (Fig. 5), as such N2 O flux estimates should be interpreted cautiously. During the first measurement period, there was

S.J. Livesley et al. / Agricultural and Forest Meteorology 151 (2011) 1440–1452

Table 3 Significant relationships between manual chamber soil CO2 and CH4 flux and soil physio-chemical properties as determined by step-wise linear regression in burnt/unburnt plots and combined data at Howard Springs, NT. There were no significant relationships for soil N2 O fluxes.

CO2 flux (ln) Burnt Unburnt All CH4 flux Burnt Unburnt All

2

Coefficients

r

p value

W (0.411), N (8.978) W (0.572), N (2.346) W (0.454), N (2.814)

0.302 0.697 0.428

0.000 0.000 0.000

nd W (3.262) A (23.621), W (2.317)

nd 0.078 0.080

nd 0.038 0.009

20

Inorganic N pool size (mg N kg-1 soil)

1446

Log normal values: W, soil %WFPS; N, soil nitrate; A, soil ammonium; T, soil temperature; nd, not determined.

NH4+ (ng N g-1 resin d-1)

NO3(ng N g-1 resin d-1)

a notable decrease in the frequency of negative N2 O fluxes during the wet season (12/07 to 03/08) (Fig. 5) which complements the seasonal pattern observed for manual chambers (Fig. 2). Methane fluxes measured by manual chambers were negative (soil CH4 uptake) and showed no significant response to burn treatment and no significant change according to season, being −12.80 ␮g C m−2 h−1 in the dry season and −10.37 ␮g C m−2 h−1 in the wet (Fig. 2). Some individual chambers measured a CH4 emission (positive flux), but mean CH4 flux for any given plot (n = 5) was invariably negative. On the first four measurement occasions (dry–wet transition and wet season), soil CH4 uptake appeared to be greater in the burnt plots than unburnt plots, however, the reverse was true in the following two dry seasons. ANOVA indicated no significant burn treatment effect (Table 2) and stepwise linear regression indicated soil CH4 flux was significantly, but weakly correlated with soil WFPS and NH4 + concentration (Table 3). As an indication of spatial variation, the %CV of manual chamber CH4 flux measurements was between 52 and 1209. Near continuous soil CH4 flux as measured by automated chambers was generally negative throughout the 15 month measurement period, indicating a net CH4 sink as seen from manual chambers (Fig. 5). Mean CH4 flux in unburnt savanna soil in the first measurement period (n = 8) was −13.1 ± 0.2 ␮g C m−2 h−1 , but ranged from −24.4 up to 23.2 ␮g C m−2 h−1 . In the second measurement period, mean CH4 flux in unburnt savanna soil was −15.4 ± 0.1 ␮g C m−2 h−1 , ranging between −24.3 and −4.5 ␮g C m−2 h−1 (Fig. 5). There were short periods of mean CH4 emission during the 2007 dry–wet transition and 2007/08 wet season because of large CH4 emissions in one or two automated chambers, for a few days up to several weeks. After the controlled burn on 27 August 2008, there was a large pulse of CH4 emission

NO unburnt NO burnt 15

NH

unburnt

NH

burnt

10

Fire 5

0 01 Sep

01 Oct

01 Nov

01 Dec

01 Jan

Fig. 3. Soil nitrate and ammonium concentrations in the five months after the controlled fire event on 27 August 2008, in both adjacent burnt and unburnt areas. Values and standard errors are from the mean of five soil separate samples (0–0.05 m).

lasting approximately 24 h, peaking at 131 ␮g C m−2 h−1 (Fig. 6). Overall, there appears to be a weak trend of greater CH4 uptake (more negative) in the dry season and less CH4 uptake (less negative) in the wet season, however the short periods of CH4 emission and movement between paired base frames mask this. Soil CO2 fluxes measured by episodic manual chambers showed a pronounced and significant seasonal pattern (Fig. 2 and Table 2) that most strongly correlated with soil WFPS (Table 3). There was no significant difference between soil CO2 flux in burnt and unburnt plots, as flux ranged from 102 to 131 mg C m−2 h−1 in the dry and dry–wet transition seasons, up to a mean flux of 238 mg C m−2 h−1 in the wet season when soil WFPS was >40%. There was less spatial variation in soil CO2 flux from manual chambers, with %CV ranging between 19 and 138 according to the time of measurement. There was a significant relationship between soil CO2 flux and soil WFPS and NO3 − concentration that was able to explain 70% of the variation in flux measured from manual chambers in the unburnt plots. The near-continuous automated chamber measurements showed large seasonal changes, but little evidence of a fire effect in response to the 27 August 2008 burn. Mean soil CO2 flux in the dry–wet transition of 2007 was 87 mg C m−2 h−1 , which increased to 108 mg C m−2 h−1 in the 2007/08 wet season (Fig. 5). In the dry season months of August and September 2008, mean soil CO2 emission was 28 mg C m−2 h−1 , which increased to 70 mg C m−2 h−1 in the dry–wet transition and 123 mg C m−2 h−1 in the December 2008 wet season month (Fig. 5). Soil CO2 emissions increased markedly in response to rainfall events in the dry–wet transition

350 250 150 50 -50 500 400 300 200 100 0 1 Oct 07

20 Dec 07

9 Mar 08

28 May 08

16 Aug 08

4 Nov 08

23 Jan 09

Fig. 4. Potential nitrate and ammonia mineralisation rates from resin bag incubation (∼0.05 m) in areas of burnt (solid line) and unburnt (dashed line) savanna woodland between October 2007 and January 2009 at Howard Springs, NT. Values are the mean of 9 resin bags, three bags in three replicate plots.

S.J. Livesley et al. / Agricultural and Forest Meteorology 151 (2011) 1440–1452

40 20

20 40 60

30

80

A

20

100 120

10

140 0 7.5

B

5.0 2.5 0.0 -2.5 -5.0 -7.5 30

-2

Unburnt savanna Burnt savanna

C

-1

CH4 flux (µg C m h )

-2

-1

N2O flux (µg N m h )

0

40

-1

Soil temperature (°C)

Soil WFPS (%)

60

Soil temperature (10 cm) Soil WFPS% (10 cm) Rain (mm)

Rainfall (mm d )

0

50 80

1447

20

131.0 peak

10 0 -10 -20

-2

-1

CO2 flux (mg C m h )

-30 250

D

200 150 100 50 0

Oct

Nov

Dec 2007

Jan 2008

Feb

Mar

Apr

May

Jun

Jul

Aug

Sep

Oct

Nov

Dec 2008

Jan 2009

Fig. 5. Soil CH4 , N2 O and CO2 exchange, soil water content, soil temperature and rainfall from October 2007 to December 2009 at Howard Springs, NT. Between October 2007 and May 2008 gas flux measurements were made in an area of unburnt savanna woodland. In June 2008 the trace gas measurement system was relocated and gas flux measurements from August 2008 to December 2009 were made in an area of unburnt savanna woodland and an area that received a controlled burn on 27 August 2008.

season (Fig. 5). As the wet season developed, the response of soil CO2 flux to individual rainfall events decreased. Soil CO2 emissions showed a far greater response to rainfall than did fluxes of N2 O and CH4 .

140

Unburnt savanna Burnt savanna

120

In controlled irrigation trials during the 2009 dry season soil water content ranged from 0.04 to 0.23 m3 m−3 or from 8.2 to 46.9% WFPS (Fig. 7). ANOVA indicated there was a significant soil moisture effect upon the soil gas diffusion coefficient (p = 0.016), soil CH4 uptake (p = 0.017) and soil CO2 emissions (p = 0.010), such that the diffusion coefficient and CH4 flux in wet soil was significantly less (LSD, p ≤ 0.05) than that in dry soil (Fig. 7), and soil CO2 flux in dry soil was significantly less (LSD, p ≤ 0.05) than that in medium and wet. In contrast, there was no significant soil moisture effect upon soil N2 O flux. 3.4. Mean daily flux (per month) and annual flux estimates Estimates of mean daily flux of CO2 , N2 O and CH4 were established for 12 consecutive months from automated chamber measurements before the controlled burn and from manual cham-

CH4 flux (µg C m-2 h-1)

3.3. Soil gas diffusion coefficients and soil moisture 100 80 60 40 20 0 -20 25/8

27/8

29/8

31/8

Date Fig. 6. Soil CH4 exchange before and immediately after a controlled savanna grass fire (27 August 2008), with concurrent measurements in an unburnt area of savanna. Positive values indicate emission and negative values soil uptake.

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S.J. Livesley et al. / Agricultural and Forest Meteorology 151 (2011) 1440–1452

Fig. 7. The effect of soil water content (A) on gas diffusivity (B), N2 O flux (C), CH4 flux (D) and CO2 flux (E) for savanna woodland soils at Howard Springs, NT. Standard errors bars presented (dry n = 8; medium n = 8; wet n = 7).

bers measurements (n = 8) in the burnt and unburnt plots from October 2007 to September 2008. For months with a direct measurement of flux, the mean daily flux was simply flux m−2 h−1 scaled up to ha−1 d−1 . For months without direct flux measurement, the mean daily flux for that month was estimated as being an average of the nearest month preceding and nearest ‘measured’ month antecedent. Annual flux was estimated from the sum of all twelve months, bearing in mind that a monthly flux was estimated as a mean daily flux multiplied by the number of days in that month. Annual flux was then converted to carbon dioxide equivalents (CO2 -e) according to their global warming potential (Solomon et al., 2007). 4. Discussion 4.1. Seasonal changes in N2 O flux In our study, soil-atmosphere exchange of N2 O occurred at very low rates throughout all seasons and soil moisture conditions (between −2.5 and 2.5 ␮g N m−2 h−1 ), which are consistent with previous savanna studies in South America (Pinto et al., 2002; Poth et al., 1995; Varella et al., 2004), North America (Martin et al., 2003; Smart et al., 1999) and Africa (Brummer et al., 2008; Castaldi and Fierro, 2005; Levine et al., 1996). In several of these studies, soil N2 O fluxes were all below MDL (Pinto et al., 2002; Smart et al., 1999; Varella et al., 2004). The small N2 O fluxes we measured is not consistent with the observation that savannas and seasonally dry

ecosystems represent a significant global source of 4.4 Tg N2 O y−1 (Castaldi et al., 2006). This is the first measurement of soil N2 O flux in Australian savanna ecosystems, and they represent ∼7% of the world’s savanna area. As such, this global emissions estimate may be an over-estimate, if savanna ecosystems in Australia and elsewhere exhibit similarly small N2 O fluxes. Our small N2 O flux measurements compliment the claim that savanna ecosystems exhibit very tight N cycling and low N availability (Bustamante et al., 2006; Pinto et al., 2002). In natural savanna ecosystems N is efficiently cycled such that little is lost through leaching or gaseous transformation (Bustamante et al., 2006). Soil N2 O is produced through either microbial nitrification or denitrification processes, and soils with small N2 O emission losses generally have a small soil inorganic N pool dominated by NH4 + (Davidson and Verchot, 2000; Firestone and Davidson, 1989) as was the case in our study, where NH4 + reached up to 20.0 mg N kg−1 whilst NO3 − remained <1.0 mg N kg−1 . Resin bags indicated that nitrification occurred in the early wet season, but did not lead to a concurrent increase in the soil NO3 − pool or soil N2 O flux. Soil N ammonification rates occurred at reasonable rates in all seasons whereas nitrification was negligible throughout the dry season, as soil microbial nitrifiers are physiologically inactive under dry soil conditions (Firestone and Davidson, 1989). Disparity between ammonification and nitrification during the dry season led to an accumulation of NH4 + , as is commonly observed in seasonally dry ecosystems (Wetselaar, 1968), which provided the substrate for nitrification observed in the early wet season.

S.J. Livesley et al. / Agricultural and Forest Meteorology 151 (2011) 1440–1452

The limited size of the soil NO3 − pool can be related to soil microbial immobilisation and preferential root uptake. Low soil N availability and mineralisation is commonly associated with large C:N ratios (>60) found in the surface litter of savanna ecosystems (Bustamante et al., 2006; Pinto et al., 2002). In Australian savannas, Russell-Smith et al. (2009) reported C:N ratios of 115 for grass litter and 123 for woody litter. Preferential uptake of NO3 is typical of annuals (e.g. sorghum) and herbs in Australian savanna ecosystems (Schmidt et al., 1998). Sorghum grass understorey establishes in the dry–wet transition producing a large N demand and fine root density capable of capturing that inorganic N. The limited accumulation of NO3 − in our resin bags suggests NO3 − leaching from large rainfall events was unlikely, but is a possibility in sandy savanna soils (Schmidt and Lamble, 2002). Some savanna studies have measured significant seasonal differences in soil N2 O flux between dry and wet seasons (Andersson et al., 2003; Castaldi et al., 2006; Davidson et al., 1993). Brummer et al. (2008) measured soil N2 O flux < 8.0 ␮g N m−2 h−1 in Burkina Faso savanna woodlands, but measured large pulse emissions up to 150 ␮g N m−2 h−1 with early rains. In our study, there was no significant change in N2 O flux between wet and dry seasons. Large, transient pulses of soil N2 O emission are often related to inorganic N accumulation through the dry season with a rapid increase in nitrifier, and possibly denitrifier, activity upon wetting (Davidson et al., 1993; Otter et al., 2001; Scholes et al., 1997). In temperate grassland systems, a WFPS of 60–65% has shown to be an approximate ‘tipping point’ between nitrification (<60%) and denitrification (>60%) dominating the production of N2 O (Müller and Sherlock, 2004; Stevens et al., 1997). At Howard Springs, WFPS only exceeded 60% for a few hours during, and immediately after, large rainfall events. The largest N2 O fluxes are generally measured in fertile savanna ecosystems with soils that have good water retention (Castaldi et al., 2006), whereas, Howard Springs has a nutrient-poor sandy soil with poor moisture retention. The uptake of N2 O by soil measured in the dry and dry–wet transition seasons has been observed in other savannas (Brummer et al., 2009; Donoso et al., 1993; Sanhueza et al., 1990). In N poor soils that experience extended dry and hot conditions, soil microbial denitrifiers may exploit N2 O as a N substrate in the absence of NO2 − and NO3 − (Donoso et al., 1993; Rosenkranz et al., 2006). Alternatively, the nitrogenase enzyme involved in biological N2 fixation in the surface soil can use N2 O as a substrate (Jensen and Burris, 1986). 4.2. Fire effects upon N2 O flux There was no apparent increase in soil NH4 + or NO3 − in response to the September 2007 controlled fire event, although data for 6 months after the fire was not available. After the 27 August 2008 fire there was no change in soil NO3 − , but soil NH4 + concentrations in the burnt sites were twofold greater than in the unburnt plots. Similarly, Pinto et al. (2002) measured no change in soil NO3 − but a twofold increase in NH4 + that persisted for 30 days following a Cerrado grassland savanna fire. At Howard Springs, the increase in soil NH4 + was no longer apparent by late-December 2008. Forest or woodland fires will generally lead to an increase in soil NO3 − and NH4 + (Attiwill and Adams, 1993), so will potentially increase N losses, such as N2 O emissions. However, whilst many savanna fire studies have measured an increase in soil inorganic N, this has not led to an increase in N2 O flux (Anderson and Poth, 1998; Andersson et al., 2003; Levine et al., 1996; Pinto et al., 2002). Savanna ecosystems generally experience a high frequency of fire, and in Northern Australia the savanna burns on average once every two years (Russell-Smith et al., 2003). As such, manipulating (prevention or promotion) fire for one or two years is insufficient to significantly change soil N mineralisation and N2 O flux, unless fire manipulation treatments are prolonged (Anderson and Poth, 1998).

1449

4.3. Seasonal changes in soil CH4 flux Soil CH4 fluxes at Howard Springs showed no obvious seasonal pattern but mean flux rates were similar to those reported in other savanna ecosystems in Africa (Castaldi et al., 2006; Otter and Scholes, 2000; Prieme and Christensen, 1999; Zepp et al., 1996), South America (Anderson and Poth, 1998) and Australian dry-temperate woodlands (Livesley et al., 2009). In Ghana, Prieme and Christensen (1999) measured mean savanna soil CH4 uptake rates of between −6.8 and −19.5 ␮g C m−2 h−1 , and observed a threefold increase in CH4 uptake after the break of rains. Low soil moisture can limit soil microbial activity, including methanotroph activity (Schnell and King, 1996). In our study there was no apparent response to the start of the wet season, and no consistent seasonal pattern in soil CH4 exchange. Zepp et al. (1996) similarly found no response in CH4 flux to early rains in South African savannas. Castaldi et al., 2004 observed a switch from soil CH4 uptake to emission from the dry season into the wet season in a temperate savanna. A critical WFPS threshold of 30% appeared to be the equilibrium point between net CH4 uptake and emission. Brummer et al. (2009) observed a switch from CH4 uptake to CH4 emission at a soil WFPS of 60–70% in Ghanaian savanna soils. At Howard Springs, soil WFPS rarely exceeded 60%, except during, or immediately after major rainfall events. Occasionally, there were short-lived CH4 emissions measured in automated chambers, and it is possible that soil methanogenesis after large rainfall events contributed. However, there are long periods when soil WFPS was high, but soil CH4 uptake was similar to that in the dry season. Anderson and Poth (1998) were able to decrease CH4 uptake by wetting up Brazilian savanna soils, as was observed in our controlled field irrigation experiment (Fig. 7). This reduction in CH4 uptake in response to increased soil moisture is principally due to reduced soil gas diffusion into and through the soil profile to the methanotrophs (Conrad, 1996; Dorr et al., 1993; King, 1997; Smith et al., 2003) as confirmed by SF6 determination of soil gas diffusion coefficients. Regardless, the relationship between soil CH4 uptake and %WFPS in manual chamber was weak, explaining only 5–6% of observed variation. This probably indicates that rapid drainage of these sandy soils limits the number of occasions gaseous diffusion can significantly limit soil CH4 uptake, i.e. methanotroph activity. The SF6 gas diffusivity measurements were made within 1 h of simulated irrigation; the episodic manual chamber measurements are unlikely to have been made within 1 h of a large natural rainfall event. This, coupled with the high spatial variability in soil bio-physio-chemical properties and processes, means it is understandable that soil %WFPS had only a weak correlatory relationship with soil CH4 uptake. Alternatively, the occasional and short-lived CH4 emissions in automated chambers may relate to termite activity. Anderson and Poth (1998) demonstrated the importance of subterranean termite activity in Brazilian savanna soils by inhibiting soil methanotroph oxidation with CH3 F which immediately changed the soil into a CH4 source. Soil CH4 emissions from subterranean termites may partially offset soil CH4 uptake (oxidation) rates, or create a net soil CH4 emission (Anderson and Poth, 1998). Termite activity beneath some automated chambers was confirmed when chambers and frames were removed. 4.4. Fire effects upon CH4 There was no significant fire effect upon CH4 flux in either the manual or automated chamber measurements, despite greater soil temperatures, less soil moisture and greater soil NH4 + concentrations. The greatest potential for soil methanotroph oxidation activity is often located between 10 and 20 cm in the soil profile (Potter et al., 1996) where disturbing fluctuations in moisture and

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S.J. Livesley et al. / Agricultural and Forest Meteorology 151 (2011) 1440–1452

Table 4 Daily mean flux (per month) and annual flux estimates for CO2 , CH4 and N2 O in savanna woodland, Howard Springs, NT. Daily mean fluxes are from near-continuous automated measurements of unburnt savanna soil, and from manual chamber measurements (n = 15) on 8 occasions in burnt and unburnt savanna plots. Positive values indicate emissions and negative values soil uptake. Daily mean fluxa

N2 O flux (g N2 O ha−1 d−1 )

CH4 flux (g CH4 ha−1 d−1 )

Automated

Automated

Automated

Manual

Manual

Manual

Unburnt

Unburnt

Burnt

Unburnt

Unburnt

Burnt

Unburnt

Unburnt

Burnt

59 102 95 94 94 105 c 65 c 45 c 35 c 30 25 25

67 119 174 262 216 c 168 119 c 89 58 c 63 c 65 c 66

98 89 203 124 277 c 228 180 c 142 105 c 102 c 100 c 99

0.04 0.09 0.08 0.19 0.12 0.11 0.08 c −0.01 c −0.05 −0.10 −0.02 −0.04

−0.10 0.02 −0.23 −0.04 0.35 c 0.13 −0.09 c −0.09 −0.09 c −0.10 c −0.10 c −0.10

−0.16 0.05 −0.13 0.10 0.43 c 0.25 0.08 c 0.07 0.06 c −0.05 c −0.11 c −0.13

−5.2 −1.5 −5.1 −4.7 −5.0 −5.8 −5.4 −6.1 −6.8 −7.5 −6.2 −5.8

−4.4 −6.1 −5.1 −1.4 −2.3 c −3.4 −4.6 c −5.3 −5.9 c −5.1 c −4.8 c −4.6

−3.7 −6.5 −6.5 −2.4 −5.9 c −6.1 −6.3 c −4.4 −2.4 c −3.1 c −3.4 c −3.5

kg ha−1 yr−1

23,498

44,385

52,876

15.07

−15.30

12.60

−1973.9

−1611.7

−1643.7

GWP kg CO2 -e ha−1 y−1

1 23,498

1 44,385

1 52,876

298 4.5

298 −4.5

298 3.7

25 −49.3

25 −40.3

25 −41.1

Octoberb November December January February March April May June July August September d e f

CO2 flux (kg CO2 ha−1 d−1 )

a b c d e f

Daily mean flux values are in kg CO2 ha−1 d−1 , g N2 O ha−1 d−1 and g CH4 ha−1 d−1 for each month. October monthly mean flux are the average of two measurements (2007 and 2008). Inferred flux from average of preceding month and antecedent measured month. Annual sum values are in kg CO2 ha−1 y−1 , g N2 O ha−1 y−1 and g CH4 ha−1 y−1 . GWP are 100 year time horizon from Solomon et al. (2007). All carbon dioxide equivalent estimates are in kg CO2 -e ha−1 y−1 .

temperature are less pronounced. Likewise, the impact of a low intensity savanna fires upon this soil methanotroph community is likely to be small. Anderson and Poth (1998) saw no significant change in CH4 uptake in response to a recent burn as compared with savanna that had not been burnt for 20 years. Similarly, Zepp et al. (1996) in South Africa saw no change in soil CH4 flux after savanna fires. In a review of savanna trace gas research, Castaldi et al. (2006) concluded that savanna fires lead to similar, or higher, soil CH4 uptake, but the mechanisms involved are unknown. Poth et al. (1995) suggested the temporary absence of termite activity after fire may lead to a net increase in soil CH4 uptake, as oxidation is no longer offset by termite emissions. The controlled fire on 27 August 2008 led to large CH4 emissions as the ash bed smouldered (Fig. 6). This is the first time such a CH4 emission has been quantified using a closed chamber technique, and it is highly likely that this is an abiotic process. Pyrovolatilisation of organic compounds have been well studied during Australian savanna fires through air-borne and ground-based flask sampling (Meyer et al., 2008).

activity thereby underpinning the significant relationship between soil nitrate and soil CO2 flux. Measurements of soil CO2 flux using automated chambers were approximately 50% of those with manual chambers, which may relate to the extended incubation period and the fact that only flux night-time measurements were reported for the automated chambers. 4.6. Fire effects upon CO2 There was no apparent impact upon soil CO2 emissions after either the 2007 or 2008 fires. Other studies have measured no significant fire effect upon soil CO2 flux in South American (Pinto et al., 2002) and South African (Zepp et al., 1996) savanna systems. The moderate-intensity of the savanna grass fires at Howard Springs suggests that soil microbial processes and root activity continue regardless. Only after prolonged fire treatments (annual burn or fire prevention) can a significant soil CO2 flux difference be expected (Pinto et al., 2002), although large CO2 emissions during the combustion event are obviously expected.

4.5. Seasonal changes in soil CO2 flux 4.7. Annual budget At Howard Springs, Chen et al. (2002) measured mean soil CO2 fluxes of 65.2 mg CO2 -C m−2 h−1 in the dry season and up to 218.6 mg CO2 -C m−2 h−1 in the wet season, producing an annual flux of 52.6 mg CO2 ha−1 y−1 . These rates are similar to our manual chamber measures of 131.2 (dry) and 237.7 (wet) mg CO2 C m−2 h−1 , and our estimate of annual soil CO2 flux from manual chamber is within 5% of that estimated by Chen et al. (2002). In our study, soil environmental variables were able to explain up to 70% of the variation in manual chamber CO2 flux, whereas Chen et al. (2002) established stronger relationships by integrating fluxes over several days and by measuring larger replicate numbers. The relationship between soil WFPS and soil CO2 flux relates to greater soil microbial and root growth activity and therefore belowground respiration as soil moisture increases after extended dry conditions (Raich and Tufekcioglu, 2000). Similarly, soil nitrate should enable greater soil microbial growth and promote greater fine root uptake

The mean annual estimates of soil CH4 exchange (Table 4) of between −40.3 and −49.3 kg CO2 -e ha−1 y−1 represent a −7.7 and −9.4 Tg CO2 -e sink if applied to Australia’s 1.9 million km2 of savanna. This biome sink estimate does not take into account CH4 emissions from areas of savanna that function as ephemeral wetlands (Deutscher et al., 2010), and does not adequately incorporate CH4 emissions from termite mounds and tree-dwelling termite colonies (MacDonald et al., 1998). From a net GHG balance perspective, this CH4 sink increases the estimated ecosystem C sink by 2%, as Beringer et al. (2007) estimated that Howard Springs had a net biome productivity (NBP) of −2000 kg CO2 ha−1 y−1 . This NBP estimate incorporates C uptake through plant growth and C loss through bi-annual grass fires. On an annual basis, the small N2 O source (or sink) estimates contribute very little to the savanna net GHG balance.

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5. Conclusion This study is the first long-term research study to quantify the seasonal variability, and to some degree inter-annual variability, of soil-atmosphere exchange of CH4 , N2 O and CO2 in an Australian savanna. A key conclusion from this study was that soil N2 O flux is very small in these N-limited savanna woodland soils in northern Australia, despite the evidence of nitrification in the early wet season period, however, there is likely to be some spatial variability in relation to vegetation structure and composition. Under certain seasonal conditions soil N2 O uptake can occur in these dry, warm N-limited savanna soils such that the net annual N2 O flux may be negative. Soil CH4 exchange in these sandy savanna soils is generally negative (uptake) apart from transient periods of soil CH4 emissions in response to either anaerobic conditions at high soil water contents or termite foraging activity. Regardless, these savanna soils are an important net CH4 sink that represents a potential −7.7 to −9.4 Tg CO2 -e sink, if applied to Australia’s 1.9 million km2 of savanna. To place this CO2 -e sink in context, this would offset 10% of Australia’s annual transport CO2 -e emissions, or almost 10% of CO2 -e emissions from Australia’s agriculture sector. Savanna grass fires had no significant long-term effect upon soilatmosphere exchange of N2 O and CH4 , however, there was a large but short-lived pulse of CH4 into the atmosphere in the 24 h after the fire combustion event. The large bimodal rainfall pattern led to strong seasonal contrast in soil moisture conditions that drove the seasonal pattern in soil CO2 flux, and contributed significantly but weakly to soil CH4 flux. The importance of termite activity and ephemeral wetlands to the CH4 balance of these globally important savanna ecosystems required targeted research. Overall, greenhouse gas exchange between the atmosphere and these savanna soils is dominated by CO2 flux, as soil-atmosphere CH4 and N2 O exchange rates are several orders of magnitude smaller. Acknowledgements This research was funded by an ARC Linkage project (LP0774812) and the Victorian Department of Sustainability and Environment. The authors would like to thank Bianca Baldissera for assistance in the field and sample processing in the laboratory and Dr. Suzanne Venn for assistance in the field. We would like to thank Andrew Edwards and Cameron Yates for assistance in the controlled (and uncontrolled) savanna burns. Similarly, thanks go to Matt Lee and Najib Ahmady of Creswick Laboratories at the University of Melbourne for their assistance in sample processing and analysis, and to Dr. Peter Isaac for providing flux tower soils data. References Anderson, I.C., Poth, M.A., 1998. Controls on fluxes of trace gases from Brazilian cerrado soils. Journal of Environmental Quality 27 (5), 1117–1124. Andersson, M., Kjoller, A., Struwe, S., 2003. Soil emissions of nitrous oxide in fire-prone African savannas. Journal of Geophysical Research-Atmospheres 108 (D20), doi:10.1029/2002JD003345. Archibald, S.A., Kirton, A., van der Merwe, M.R., Scholes, R.J., Williams, C.A., Hanan, N., 2009. Drivers of inter-annual variability in Net Ecosystem Exchange in a semi-arid savanna ecosystem. South Africa. Biogeosciences 6 (2), 251–266. Attiwill, P.M., Adams, M.A., 1993. Nutrient cycling in Forests. New Phytologist 124 (4), 561–582. Beringer, J., Hutley, L.B., Tapper, N.J., Cernusak, L.A., 2007. Savanna fires and their impact on net ecosystem productivity in North Australia. Global Change Biology 13 (5), 990–1004. Beringer, J., Hutley, L.B., Tapper, N.J., Coutts, A., Kerley, A., O’Grady, A.P., 2003. Fire impacts on surface heat, moisture and carbon fluxes froma tropical savanna in northern Australia. International Journal ofWildland Fire 12 (3–4), 333–340. Beringer, J., Hutley, L.B., Hacker, J.M., Neininger, B., Paw, U.K.T., 2011. Patterns and processes of carbon, water and energy cycles across northern Australian landscapes: From point to region, Introduction to Special issue. Agricultural and Forest Meteorology. (under review). Bousquet, P., Ciais, P., Miller, J.B., Dlugokencky, E.J., Hauglustaine, D.A., Prigent, C., Van der Werf, G.R., Peylin, P., Brunke, E.G., Carouge, C., Langenfelds, R.L., Lath-

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