Introduction

The degree to which CH₃Br released in the lower atmosphere migrates to the stratosphere and releases Br to catalyze ozone destruction is directly proportional to its tropospheric lifetime [Solomon et al., 1992; Mellouki et al., 1992; Kurylo and Rodriguez, 1999]. Its lifetime is influenced by the rate at which CH₃Br reacts with hydroxyl radicals in the troposphere, the rate at which it is absorbed into surface waters, and the rate at which it is taken up into and destroyed in soil and/or on vegetation surfaces. The first loss term (reaction with hydroxyl radical) is reasonably well known [Mellouki et al., 1992]. The second (oceanic uptake) is estimated from CH₃Br solubility, diffusivity, and surface exchange coefficient [Butler, 1994]. There are few data to allow an estimate of the third term (soil consumption). It had been observed that CH₃Br could be consumed in agricultural soils and this consumption was linked to ammonium fertilizer applications [e.g., Ou et al., 1994]. It was also demonstrated that anaerobic soils when exposed to high levels could also consume CH₃Br [Oremland et al., 1994]. However it was only recently demonstrated by the work of Shorter et al. [1995] that a wide variety of aerobic, drained, near-surface upland soils consume CH₃Br at near-ambient tropospheric mixing ratios. This observation has been subsequently corroborated by others [e.g., Serça et al., 1998]; it appears to be a ubiquitous process of aerobic terrestrial soils.

The discovery of a hitherto unrecognized sink had two effects. One, it put the tropospheric budget further out of balance than it had been previously and, two, it lowered the ozone depletion potential from 0.6 to 0.4 [Kurylo and Rodriguez, 1999]. In terms of the balance between sources and sinks, either the tropospheric lifetime of CH₃Br had to be seriously overestimated (unlikely) and/or there had to be previously unidentified sources.

The soil sink of tropospheric CH₃Br

The principal sinks of atmospheric CH₃Br include reaction with OH [Mellouki et al., 1992], irreversible loss to the ocean [Butler, 1994; Yvon and Butler, 1996], uptake by soils [Shorter et al., 1995; Serça et al., 1998], and possibly uptake by green plants [Jeffers et al., 1998]. The sinks total approximately 205 Gg/yr [Kurylo and Rodrigues, 1999; Yvon-Lewis and Butler, 1997; see article by Yvon-Lewis in this issue for the most recent modifications], with the earliest reports of the soil sink contributing between 20% (42 ± 32 Gg y^-1, [Shorter et al., 1995]) and 45% (94 ± 54 Gg y^-1, [Serça et al., 1998]) of the total.

The Shorter et al. [1995] report was the first attempt to estimate the global uptake of ambient CH₃Br by soils. The research involved measuring the uptake of CH₃Br by selected soils from a New Hampshire forest, cornfield, and grassy field, as well as soils from Costa Rican tropical forest and Canadian boreal forest. Consumption into surface soils from enclosed headspace [chamber fluxes] was also observed. In the direct chamber measurements the loss rate of CH₃Br was greater than the loss of an inert tracer (SF6) [Shorter et al., 1995, Serça et al., 1998] indicating an active consumption process. The rate of chamber-measured influx was consistent with flux rates extrapolated from soil incubation techniques. The use of a soil incubation assay for consumption allowed broad surveys to be carried out relatively quickly. Since the Shorter et al. [1995] paper, more than 230 soil samples have been collected from over 90 sampling locations across the United States, Alaska, Canada, Costa Rica, Brazil, Germany, China, Finland and Siberia. The soils were collected at agricultural, forest, meadow, pasture and desert locations, representing most of the major soil classification groupings. Consumption of CH₃Br could be measured in all the soil samples (see Table 1).

Table 1: Rates of uptake of CH3Br by soils from different biomes.

1Shorter et al.. 1994; 2Varner et al.. 1999a; 3Serça et al.. 1998

The temperate zone soils had the most rapid uptake of CH₃Br while the northern and tropical zones were consistently less active by about 10%. Surface soils were on average approximately 50% more active than deeper soils. In a few isolated cases deeper agricultural soils were more active. Similar rates of uptake of ambient CH₃Br were observed with direct chamber measurements of drained upland forest and agricultural soils in New Hampshire [Varner and Crill, unpublished data]. In addition Serça et al. [1998] reported bog microcosms to consume 2.1 +/- 0.9 Gg y-1. However, direct chamber flux measurements indicate natural wetlands are net sources of CH₃Br rather than net sinks (see below).

Green plant material from a wide variety of species has also been shown in laboratory incubations of tissue under ppmv levels to be capable of consuming CH₃Br [Jeffers and Wolfe, 1997; Jeffers et al., 1998]. Even though the authors point out that their observations are consistent with an enzymatic degradation, it is still uncertain how this fits into the atmospheric budget or whether this mechanism is operative under ambient (part-per-trillion by volume) mixing ratios. The role of plants in the consumption (and production) of atmospheric CH₃Br remains an exciting avenue of investigation.

The use of methyl halides by bacteria has been studied (but not in great detail) for more than 20 years. In the case of CH₃Br, more studies have concentrated on aerobic consumption because of fears of release of CH₃Br from soils during agricultural fumigation practices [Yates et al., 1996a,b; Miller et al., 1997]. Methyl halides can be consumed by whole cells and cell extracts of methanotrophic bacteria [Oremland et al., 1994, Goodwin et al., 1998]. Nitrifying bacteria also consume methyl halides via ammonia monooxygenase [Rasche et al., 1990], and additions of ammonia fertilizers stimulate consumption by agricultural soils [Ou et al., 1997].

Methyl halides can also be consumed microbially in anaerobic environments. CH₃Br can be utilized through a nucleophilic substitution with sulfide, which produces methylated S gases [Oremland et al., 1994b].

The above studies were performed with mixing ratios or concentrations of substrate that were orders of magnitude larger than ambient levels. Since then, techniques have been developed that allow us to measure the process of CH₃Br consumption at near-ambient levels. We have learned the following from these studies [cf., Hines et al., 1998]:

• Consumption in aerobic soils is biologically mediated. The temperature response of the activity-specifically, declining activity above 30oC-is consistent with biological processes. Autoclaving stops consumption altogether.
• Consumption is bacterial. Consumption is sensitive to antibiotics specific to bacteria but not sensitive to inhibitors that affect higher organisms.
• Rates of consumption appear to be sensitive to soil moisture and/or soil organic matter (SOM) content. Drier soils and those poor in organic content demonstrate slower rates. SOM and soil moisture are not independent variables, so it is difficult to distinguish the effect of either.
• The activity is aerobic. Flooding soil samples with nitrogen slows or stops the activity.
• The depth distribution of the activity in soils is different from that of methane oxidation. Unlike methane oxidation, where the activity is greatest a few centimeters deep in the soil just at the base of the surface organic layer, CH₃Br consumption activity is always greatest at the surface except in the driest agricultural soils.

Non-anthropogenic terrestrial sources of tropospheric CH₃Br

The few estimates of global sources of CH₃Br per annum are listed by Yvon-Lewis and Butler [1997] and in the latest Scientific Assessment of Ozone Depletion [Kurylo and Rodrigues, 1999; see also article by Yvon-Lewis in this issue.] The only number known with any certainty is the annual production by the CH₃Br industry of 66 Gg. Of that amount, 20-80% is thought to make its way into the troposphere [Yagi et al., 1995]. The oceanic source of 56 Gg y^-1 is exceeded by its sink of 77 Gg y^-1, leaving a small net flux from the atmosphere to the ocean. Expanding the work of Harper et al. [1986], Lee and Holland [1999] suggest that the ability of Phellinus spp. fungi to methylate halogens could result in a potential source of 0.5 to 5.2 Gg y-1 to the global atmosphere from forest soils.

Biomass burning is another important source of tropospheric methyl halides [Manö and Andreae, 1994, Blake et al., 1996]. Biomass burning may inject into the troposphere an amount of CH₃Br that is similar in magnitude to oceanic and industrial sources. The uncertainty in the estimates is such that burning could contribute about 20% of the total source.

As noted above, the global budget of CH₃Br has been significantly out of balance with known sinks greater than identified sources by over 80 Gg y-1. Recently, this number has been reduced to 60 Gg y-1, owing to the discovery of specific plant sources. Within the past two years, the field observations of three graduate students, Robert Rhew of Scripps Institution of Oceanography (UC San Diego), Claudia Dimmer of University of Bristol, UK, and Ruth Varner at the University of New Hampshire, have revealed potentially important new sources of CH₃Br (and other methylated halogens) from salt marshes and freshwater wetlands. Also, Kelly Redecker at UC-Irvine has been working to characterize the emissions of CH₃Br from the anthropogenic wetlands of rice agriculture.

Laboratory incubations of small amounts of material collected in the field have not identified peats or other organic rich soils to be sources of CH₃Br. This is in part due to the small sample sizes and the attendant analytical difficulties. Direct field measurements use large chambers or boxes (often temperature controlled and mixed with fans) to isolate an atmosphere over a soil surface, and changes in the mixing ratios of CH₃Br are measured by sequentially sampling the enclosed headspace. A chemically and biologically inert tracer (SF6 or a CFC) is sometimes used to measure the dilution and the diffusive and advective loss rates of the gas from the chamber.

Dimmer et al. [1999] and Varner et al. [1999b] have measured effluxes directly in Irish peatlands and in New Hampshire wetlands, respectively; extrapolation yields net annual global emissions of CH₃Br of about 5.0 and 4.6 Gg. The extrapolated fluxes from rice fields are significantly less, at 1.5 Gg [Redecker et al., 1998]. Rhew et al. [1999] have observed substantially higher emissions from coastal salt marshes that extrapolate globally to 14 Gg y-1. As they point out, this would mean that 0.1% of the global surface area could produce roughly 10% of the total flux. If 60 Gg y-1 are from fumigation losses and leaded gasoline burning and all of the biomass burning is considered non-anthropogenic, then the natural emission is 80 Gg y-1 and the contribution of salt marshes would be closer to 18% of the total.

New sources of atmospheric CH₃Br have been discovered and the relative magnitudes of those sources are just beginning to be quantified. We need to understand both the spatial and temporal variability in the emissions across a landscape and why the variability is so great. The accumulation and flux of methyl halides in wetlands must be controlled by both production and consumption processes, and these processes are probably controlled by biogeochemical conditions.

Plants, specifically those from the Brassicaceae family, have been shown to produce CH₃Br when grown in soils with elevated bromide concentrations [Gan et al.,1998]. Leaf disks from a number of different plants of different species and different families will produce methyl halogens when floated in solutions with elevated halide concentrations [Saini et al., 1995]. Gan et al. suggest that one crop alone (rapeseed) could be contributing a net of 7 Gg y-1 to the atmospheric budget. Saini et al. [1995] point to the likely involvement of a non-specific methyl transferase. This would also imply a correlation between CH₃Br production and that of other halogenated methanes, such as methyl chloride, which has been observed in the flux studies [Rhew et al., 1999; Varner et al., 1999b]. Rhew et al. [1999] also noticed a correlation between green plant density and emission implicating plants in the production and/or transport of CH₃Br from salt marsh soils to the atmosphere.

There has been considerable progress recently in quantifying and understanding the variability of CH₃Br exchange with the atmosphere. Our beginnings remind us of how much remains to be done to address these important issues. Tasks include the following:

• Determine the rate and ubiquity of CH₃Br (and other halogenated methanes when possible) exchange from direct measurement surveys across aerobic soil types from different biomes, along trophic gradients of freshwater wetlands, and across salinity gradients of coastal salt marshes.
• Determine quantitatively the controls and the biological dynamics of formation and consumption of CH₃Br (and other halogenated methanes when possible) with laboratory studies and field measurement programs designed to resolve spatial variability with specific biome types.
• Determine the magnitude and the seasonality of exchange rates of CH₃Br at a number of field sites by frequent direct measurements of soil exchange.
• Develop techniques to resolve quantitatively the simultaneous processes of consumption and production (the gross fluxes) of methyl bromide from terrestrial systems, by use of isotopes or other novel approaches. This understanding is needed to obtain an accurate estimate of the partial atmospheric lifetime with respect to these sinks, and to evaluate or predict the effects of global change upon the emission of CH₃Br and, ultimately, its atmospheric burden.

Acknowledgements

The author would like to thank C. Dimmer, K. Redecker, R. Rhew, R. Varner, R. Weiss, J. Butler and R. Talbot for their kind collegial comments and contributions to this article.

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