Abstract

The oxidation of methane by methane-oxidising microorganisms is an important link in the global methane budget. Oxic soils are a net sink while wetland soils are a net source of atmospheric methane. It has generally been accepted that the consumption of methane in upland as well as lowland systems is inhibited by nitrogenous fertiliser additions. Hence, mineral nitrogen (i.e. ammonium/nitrate) has conceptually been treated as a component with the potential to enhance emission of methane from soils and sediments to the atmosphere, and results from numerous studies have been interpreted as such. Recently, ammonium-based fertilisation was demonstrated to stimulate methane consumption in rice paddies. Growth and activity of methane-consuming bacteria in microcosms as well as in natural rice paddies was N limited. Analysing the available literature revealed that indications for N limitation of methane consumption have been reported in a variety of lowland soils, upland soils, and sediments. Obviously, depriving methane-oxidising bacteria of a suitable source of N hampers their growth and activity. However, an almost instantaneous link between the presence of mineral nitrogen (i.e. ammonium, nitrate) and methane-oxidising activity, as found in rice soils and culture experiments, requires an alternative explanation. We propose that switching from mineral N assimilation to the fixation of molecular nitrogen may explain this phenomenon. However, there is as yet no experimental evidence for any mechanism of instantaneous stimulation, since most studies have assumed that nitrogenous fertiliser is inhibitory of methane oxidation in soils and have focused only on this aspect. Nitrogen as essential factor on the sink side of the global methane budget has been neglected, leading to erroneous interpretation of methane emission dynamics, especially from wetland environments. The purpose of this minireview is to summarise and balance the data on the regulatory role of nitrogen in the consumption of methane by soils and sediments, and thereby stimulate the scientific community to embark on experiments to close the existing gap in knowledge.

Introduction

Despite its low atmospheric mixing ratio (1.7 μl l−1) and short atmospheric residence time (about 10 years), CH4 is considered as the most potent greenhouse gas after carbon dioxide [1]. This is due to the higher effectiveness (20–30 times) at absorbing long-wave radiation in comparison to CO2, and due to the involvement of CH4 in chemical reactions leading to the formation of ozone [2]. Because CH4 concentration in the atmosphere has more than doubled in the post-industrial era, much of the research efforts have been expended to identify sources and sinks of methane, and to estimate their strengths. Methane flux measurements have demonstrated that soils are the most important biological sources and sinks of atmospheric methane (cf. [3]). The balance between the production of methane by methanogenic bacteria under anoxic conditions and the consumption of methane by methanotrophic bacteria under oxic conditions determines whether a particular soil is a net source or a sink of atmospheric methane.

Submerged wetland soils (e.g. swamps, bogs, rice paddies) are regarded as the most important source of atmospheric methane. The contribution of these ecosystems to the annual global methane emission has been estimated at 55%[3]. Non-flooded upland soils (e.g. forests, grassland, arable) are regarded as the only biological sink of atmospheric methane and are responsible for 6% of the global methane consumption [3]. In both wetland and upland soils, obligate aerobic methanotrophic bacteria use molecular oxygen to oxidise methane to CO2 and cell carbon [4]. In wetland soils these bacteria are active in the surface layers and in the rhizosphere of oxygen-releasing plants (cf. [5]), where they substantially reduce the potential amount of emitted methane [5]. Factors that limit or even inhibit the activities of methanotrophic bacteria have major effects on the global methane budget. Since the finding that nitrogenous fertilisation represses methane consumption in forest soils [6], more than a decade of research has focused on elucidating this aspect (cf. [3,7]). However, recently it was found that ammonium-based fertilisation stimulated growth and activity of methane oxidisers in the rhizosphere of rice [8,9]. Moreover, upon depletion of mineral nitrogen in natural rice paddies, methane oxidation decreased to zero (cf. [10]). A restudy of literature revealed a high number of articles already in 1993 and onwards showing the same effect and providing data on the importance of mineral nitrogen for the consumption of methane in soils. However, their implications have never been included in the discussions about global methane budgets. This minireview will summarise the results from these early and from the recent studies, in order to firmly establish that mineral nitrogen can be one of the limiting factors for growth of methanotrophic bacteria in soils and sediments and can even be a prerequisite for methane-oxidising enzyme activity in these ecosystems.

Nitrogen as an inhibiting factor of methane consumption in soils: ‘one side of the coin’

To fully comprehend the potentially inhibitory effects of nitrogenous fertilisers on methane oxidation we have to bear in mind that two types of kinetics have been encountered with respect to the consumption of methane in soils and sediments. The first kinetic pattern of methane oxidation, known as ‘low-affinity’ methane consumption, is observed in all methane-producing soils (e.g. wetland, peat, landfill) and is carried out by ‘conventional’ type I and II methanotrophs that display Km values in the μM range.

The second type, known as ‘high-affinity’ methane oxidation [11] occurs in soils that receive methane only by diffusion from the atmosphere (e.g. forest soil). These soils have methane concentrations in the nM range and display low apparent Km values for methane uptake. Since the apparent Km values of the monooxygenase enzyme systems of all cultured bacteria able to oxidise methane (methanotrophic bacteria and nitrifiers) are an order of magnitude higher, as yet uncultured organisms with novel variants of methane monooxygenase (MMO) are believed to be responsible for the high-affinity consumption. One study demonstrates that a normal Methylocystis strain can display high-affinity activity under certain conditions, and that the above assumption needs not to be true [12]. However, molecular data in combination with isotope and radiotracer studies have indicated that bacteria responsible for the process of atmospheric methane consumption in many soils are indeed taxonomically novel and only distantly related to the known type II methanotrophic bacteria [13–15].

The first report on inhibition of methane oxidation by nitrogenous fertilisers came from soils displaying high-affinity oxidation. In 1989 Steudler and co-workers [6] conducted a study that elucidated factors controlling biological sinks of methane. Their aim was to assess the causes of increasing atmospheric methane concentrations. One of the factors they investigated was the application of nitrogenous fertilisers to mimic the effect of the increasing atmospheric nitrogen deposition in industrialised countries on methane consumption by soils. Application of NH4NO3 to acid forest soils reduced the uptake of atmospheric methane by these soils for up to 33%. The potential implication of this observation for the global methane budget initiated numerous studies assessing the effects of nitrogenous fertilisers on methane consumption in various soils. For a detailed discussion on the literature addressing this aspect, some excellent papers can be consulted [3,7,16,17]. From the studies reviewed in these papers it is evident that the observation of Steudler and co-workers extended far beyond forest soils. Methane consumption in both upland and wetland soils is affected by nitrogenous input, although this generally holds only for ammonium-based additions. Nitrate has been found inhibitory only in very high concentrations, which likely give rise to osmotic effects. Therefore, in this article we will focus on the effect of ammonium only. In Table 1 the results of 53 studies are presented with regard to the type of habitat, the nature of the inhibitory effect, the proposed mechanisms of inhibition and whether the experiments were performed in situ or in vitro. The diversity of effects, the type of habitats, and the number of proposed underlying mechanisms indicate that no generalisations can be made with respect to the effect of ammonium-based nitrogen input on methane oxidation in different soil or sediment ecosystems. The proposed mechanisms operate at the cellular level, community level and at the level of the ecosystem. Immediate, short-term inhibition in soil or slurry incubations can be explained fairly well by the effects at the cellular level, such as competitive inhibition of MMO by ammonia (cf. [33]). Besides the oxidation of methane, the MMO also has the ability to convert ammonia to nitrite, and ammonia will therefore reduce the amount of methane consumed by methanotrophic bacteria in a concentration-dependent way. The intermediates and end products of methanotrophic ammonia oxidation, i.e. hydroxylamine and nitrite, can be toxic to methanotrophic bacteria and will also lead to inhibition of methane consumption [74]. Finally, the addition of high amounts of ammonium salts to laboratory incubations may also affect methane oxidation due to osmotic stress [22].

1

Inhibition patterns of methane oxidation following nitrogen-based fertilisation or atmospheric deposition in various soil/sediment habitats and hypothesised underlying mechanisms

Nature of the effect Habitat References reporting on in situ or in vitro inhibition of methane oxidation Proposed mechanisms of inhibition as reported in literature 
  In situ In vitro  
Short-term Forest soils (boreal as well as temperate) [18–20[21,18,16,22–26Full/partial competitive inhibition of the MMO. 
 Grassland soils  [27Nitrite/hydroxylamine toxicity. 
 Arable soils  [28–30Osmotic effects due to salt additions. 
 Agricultural soils [31,32[21,32,33 
 Peat [34[34 
 Tropical pasture soils [35  
 Landfill cover soils  [36–39 
 Wetland sediments [40,41[21,40,42–44 
 Lake sediment [45[45,46 
     
Delayed Agricultural soils [47 Initial protection from ammonium by soil N dynamics (e.g. immobilisation, nitrification). 
 Forest soils [48 Adsorption of ammonium to soil matrix (cation exchange capacity, CEC). 
    Population shift in the MOB community. 
    Inhibition of cell growth and de novo enzyme synthesis leading to delayed decrease of methane oxidation following natural cell mortality. 
     
Long-term Forest soils [6,49–53[54Damage to MOB due to osmotic stress. 
 Agricultural soils [31,52[31Damage due to exposure to nitrite. 
 Arable soils [55 Cell death due to starvation (NADH limitation). 
 Grassland soils [52,56[57Shifts in the MOB community. 
 Alpine meadows [58  
     
No effect Forest soils (boreal and temperate) [19,48,50,59–62[21,23,43,63Plant uptake of added ammonium. 
 Grassland soils [64 MOB active in subsurface layers not reached by the fertiliser. 
 Agricultural soils [65,66 Ammonium-tolerant population. 
 Arable soil [31,67–69[29Increase in ammonia oxidisers which oxidise methane. 
 Boreal peat soils [29,70[71Soil moisture status may vary and subsequent methane diffusion limitation masks inhibition by ammonium. 
 Wetland soils [72[73 
Nature of the effect Habitat References reporting on in situ or in vitro inhibition of methane oxidation Proposed mechanisms of inhibition as reported in literature 
  In situ In vitro  
Short-term Forest soils (boreal as well as temperate) [18–20[21,18,16,22–26Full/partial competitive inhibition of the MMO. 
 Grassland soils  [27Nitrite/hydroxylamine toxicity. 
 Arable soils  [28–30Osmotic effects due to salt additions. 
 Agricultural soils [31,32[21,32,33 
 Peat [34[34 
 Tropical pasture soils [35  
 Landfill cover soils  [36–39 
 Wetland sediments [40,41[21,40,42–44 
 Lake sediment [45[45,46 
     
Delayed Agricultural soils [47 Initial protection from ammonium by soil N dynamics (e.g. immobilisation, nitrification). 
 Forest soils [48 Adsorption of ammonium to soil matrix (cation exchange capacity, CEC). 
    Population shift in the MOB community. 
    Inhibition of cell growth and de novo enzyme synthesis leading to delayed decrease of methane oxidation following natural cell mortality. 
     
Long-term Forest soils [6,49–53[54Damage to MOB due to osmotic stress. 
 Agricultural soils [31,52[31Damage due to exposure to nitrite. 
 Arable soils [55 Cell death due to starvation (NADH limitation). 
 Grassland soils [52,56[57Shifts in the MOB community. 
 Alpine meadows [58  
     
No effect Forest soils (boreal and temperate) [19,48,50,59–62[21,23,43,63Plant uptake of added ammonium. 
 Grassland soils [64 MOB active in subsurface layers not reached by the fertiliser. 
 Agricultural soils [65,66 Ammonium-tolerant population. 
 Arable soil [31,67–69[29Increase in ammonia oxidisers which oxidise methane. 
 Boreal peat soils [29,70[71Soil moisture status may vary and subsequent methane diffusion limitation masks inhibition by ammonium. 
 Wetland soils [72[73 

The other inhibition patterns (delayed, long-term and no effect) act at the community or ecosystem level and are more difficult to explain. The only mechanism explaining all three patterns would be the changes of the community composition, either by a shift between ammonium-tolerant and ammonium-intolerant methane-oxidising species, or by a relative increase of ammonia oxidisers consuming methane. Generating experimental proof for these hypotheses, taking into account rapid methodological progress in molecular microbial ecology, will be one of the challenges for the researchers in near future. The prerequisites (e.g. presence of plants, cation exchange capacity, soil N dynamics, soil water status, uncoupling of population growth and activity, spatial arrangement of populations in the soil profile) for any of the other proposed mechanisms listed in Table 1 may vary considerably between soil ecosystems. Detailed knowledge on soil physicochemical parameters and on the type of methane oxidisers present in any particular environment is necessary to understand the modes of action of fertiliser application. Also, a consideration of the methodological assessment of fertiliser effects is advisable. Table 1 clearly shows that laboratory incubations (in vitro) mostly result in immediate short-term effects in upland as well as lowland soils, while the majority of the studies performing in situ flux analyses reveal long-term effects or no effect. These discrepancies strongly suggest that a comprehensive investigation of fertiliser effects on methane oxidation should combine in vitro and in situ methodologies.

Nevertheless, 15 years after the findings of Steudler and co-workers, it is evident that their results cannot simply be extrapolated to various soil ecosystems. Without any doubt nitrogenous fertiliser additions can affect methane consumption, but this is not relevant for every soil or sediment ecosystem. Before starting new studies to assess the impact of ammonium-based N input on methane oxidation in soils and to unravel the underlying mechanism, it is advisable to consider whether N-based inhibition is expected to occur in the ecosystem under study. Forest soils with high deposition of atmospheric nitrogen or agricultural soils with high fertiliser input seem to be the most relevant systems for further study in this respect. In pristine, natural soils and also sediments, the role of nitrogen as a limiting factor for methane oxidation may be of more environmental and ecological importance.

Nitrogen as a stimulating factor of methane oxidation in soils: ‘the other side of the coin’

Although the mechanisms of inhibition are still under debate, the studies supporting fertiliser nitrogen as a detrimental factor of methane oxidation are numerous. Potentially large environmental consequences of this trend have led to a rather one-sided research approach with respect to the regulatory role of nitrogen on methane consumption in soils and sediments. Recently, results have been obtained that put this relationship into a different perspective. Bodelier and co-workers [8,9] subjected either planted or unplanted rice soil microcosms to different fertiliser regimes. The activity and the population size of the methanotrophic bacteria in the rice rhizosphere were substantially enhanced by the addition of urea or (NH4)2PO4, as displayed in Fig. 1. Ammonium in the rhizosphere of unfertilised plants was depleted within 30 days of a total incubation period of 84 days. The absence of mineral nitrogen resulted in an inactive, and probably non-growing methanotrophic community. Using the same microcosm system, Eller and Frenzel [75] showed that in situ rhizospheric methane oxidation, determined by the use of specific inhibitor CH2F2, decreased to zero upon depletion of ammonium in the soil. Identical results were found in natural rice paddies [10,76]. Addition of fertiliser nitrogen to these natural rice paddies also led to the stimulation of in vitro and in situ methane consumption [10,76,77].

1

Methane-oxidising activities (A), most probable numbers (B) and PLFA (C) abundance in rice soil microcosms as determined 68 days after planting. Displayed are the effects of ammonium-based fertilisation and rice plants on the methane-oxidising community in unplanted and in rhizosphere soil from microcosms that were either unfertilised or supplemented with urea (200 or 400 kg ha−1) or (NH4)2PO4 (200 or 400 kg ha−1). The bars in A represent the arithmetic means of initial (black bars) and induced (grey bars) oxidation rates of four replicate microcosms. The bars in C represent the total abundances of type I specific PLFA (C16:1w8c) (grey bar) and type II specific PLFA (C18:1w8c) (black bar). Data displayed in A and B were analysed using the Kruskal–Wallis test. This test revealed significant (P<0.05) effects of the presence of roots and fertiliser application on methane-oxidising activities (A) as well as on numbers (B).

1

Methane-oxidising activities (A), most probable numbers (B) and PLFA (C) abundance in rice soil microcosms as determined 68 days after planting. Displayed are the effects of ammonium-based fertilisation and rice plants on the methane-oxidising community in unplanted and in rhizosphere soil from microcosms that were either unfertilised or supplemented with urea (200 or 400 kg ha−1) or (NH4)2PO4 (200 or 400 kg ha−1). The bars in A represent the arithmetic means of initial (black bars) and induced (grey bars) oxidation rates of four replicate microcosms. The bars in C represent the total abundances of type I specific PLFA (C16:1w8c) (grey bar) and type II specific PLFA (C18:1w8c) (black bar). Data displayed in A and B were analysed using the Kruskal–Wallis test. This test revealed significant (P<0.05) effects of the presence of roots and fertiliser application on methane-oxidising activities (A) as well as on numbers (B).

In order to assess whether this effect was confined to rice paddies, we examined the literature for similar findings in other environments. Table 2 lists studies that measured direct stimulatory effects of ammonium- or nitrate-based fertilisation on methane consumption. The table also lists studies that demonstrated positive correlations between soil mineral nitrogen and methane consumption rates. The latter studies were taken into account because the causal mechanisms may very well be the same as in the studies where fertiliser was actually added. The range of soils in which nitrogen availability seems to be a limiting factor for methane oxidation is as wide as for those soils where inhibition is observed. Consumption of atmospheric methane (i.e. high-affinity methane oxidation) as well as the oxidation of elevated methane concentrations (i.e. low-affinity methane oxidation) can be enhanced by both the addition of ammonium or nitrate.

2

Synthesis of studies that demonstrated positive effects of soil/sediment mineral N content and or fertilisation on methane Oxidation

Habitat High/low CH4 Fertiliser Effect Explanation proposed by the authors Reference 
Surface layer littoral sediment high NH4Cl Positive relation between NH4 and VmaxArtefacts. [45
Spruce forest soil low  High CH4 consumption with highest NH4 in soil. None. [78
Native and invaded heath land low mix NPK, 56 and 112 kg ha−1 Positive relationship between ammonium concentration and CH4 oxidation in seven heath land sites. Indirect effect of increased grass vegetation which results in higher nutrient availability. [79
Forest soils, urban to rural gradient low none Positive correlation between NH4+ in soil and methane uptake rates in acidic forest soil. Reduced nutrient availability (N, P) due to slow organic matter degradation in urban areas. [80
Meadow cambisol, cultivated cambisol, forest luvisol and paddy soil high and low (NH4)SO4 Shortening of induction times and stimulation of Vmax in short-term assays. Methanotrophs need nitrogen as an N source. [81
Grassland and deciduous forest soil low none Higher methane oxidation in forest soils with high ammonium and nitrate concentrations. Higher nitrogen turnover in grassland soils leads to higher nitrification and therefore inhibition of methane oxidation. [82
Arable and undisturbed woodland and grassland soil low none Higher oxidation rates and numbers in undisturbed woodland and grassland as compared to unfertilised arable soil. None. [83
Arable, forest and set aside soil low none Positive correlation between ammonium and methane oxidation; higher oxidation in fields taken out of production longer. None. [84
Deciduous and spruce forest soil low KNO3 or (NH4)SO4 in vitro Stimulation of methane oxidation in spruce forest due to fertilisation. General improvement of living conditions by narrowing the C/N ratio. [85
Landfill soil high NH4Cl and KNO3 Lower methane oxidation when N becomes limiting; addition of ammonium and nitrate stimulated methane consumption. N limitation. [86
Landfill soil high NH4Cl, KNO3, lime Rapid onset of methane oxidation with ammonium and nitrate in soil columns; nitrate also stimulated in fresh landfill soil. None. [36
Coniferous forest soil low CaNH4NO3 Higher methane oxidation in forest soils which have been previously fertilised (2 years). None. [87
Rice field soil high NH4Cl Inhibition turns into stimulation by ammonium after several incubations with different methane mixing ratios. Nitrifiers are stimulated and are therefore consuming more methane. [43
Landfill cover soil high NH4Cl, nitrified sludge, compost Stimulation of methane oxidation by ammonium of soil exposed to high methane for short period; inhibition increases with exposure time to high methane. Change in the community composition to N-susceptible methanotrophs. [37
Tropical pasture soil low (NH4)2SO4, urea, CaNO3 Fertilised pastures had higher methane uptake than traditional and legume pastures. Higher plant biomass in fertilised pastures results in lower water filled pore space and hence higher diffusion of methane into the soil. [35
Rice field soil high urea Stimulation of in vitro methane oxidation. N-limitation of MOB. [77
Natural (forest/prairie) and agricultural soils low various regimes Positive correlation between mineral N of soil and methane consumption at elevated levels. Increasing contribution of nitrifiers to methane oxidation. [88
N-limited forest soil low (NH4)2SO4 Higher methane uptake in N-limited soils after initial short-term inhibition. N limitation of the methanotrophs. [20
Landfill cover soil high NH4Cl, (NH4)2SO4 Stimulation of methane oxidation at high methane (2%) only when sufficient N was present. Nitrogen limitation of type I methanotrophs. [38
Agricultural soil low various regimes High levels of fertilisation (150 kg N ha−1 year−1) resulted in higher methane uptake of cultivated soils. Higher growth of methanotrophs; fertilised plots had not been tilled for a long period leading to different soil pore structure and better methane diffusion into the soil. [89
Habitat High/low CH4 Fertiliser Effect Explanation proposed by the authors Reference 
Surface layer littoral sediment high NH4Cl Positive relation between NH4 and VmaxArtefacts. [45
Spruce forest soil low  High CH4 consumption with highest NH4 in soil. None. [78
Native and invaded heath land low mix NPK, 56 and 112 kg ha−1 Positive relationship between ammonium concentration and CH4 oxidation in seven heath land sites. Indirect effect of increased grass vegetation which results in higher nutrient availability. [79
Forest soils, urban to rural gradient low none Positive correlation between NH4+ in soil and methane uptake rates in acidic forest soil. Reduced nutrient availability (N, P) due to slow organic matter degradation in urban areas. [80
Meadow cambisol, cultivated cambisol, forest luvisol and paddy soil high and low (NH4)SO4 Shortening of induction times and stimulation of Vmax in short-term assays. Methanotrophs need nitrogen as an N source. [81
Grassland and deciduous forest soil low none Higher methane oxidation in forest soils with high ammonium and nitrate concentrations. Higher nitrogen turnover in grassland soils leads to higher nitrification and therefore inhibition of methane oxidation. [82
Arable and undisturbed woodland and grassland soil low none Higher oxidation rates and numbers in undisturbed woodland and grassland as compared to unfertilised arable soil. None. [83
Arable, forest and set aside soil low none Positive correlation between ammonium and methane oxidation; higher oxidation in fields taken out of production longer. None. [84
Deciduous and spruce forest soil low KNO3 or (NH4)SO4 in vitro Stimulation of methane oxidation in spruce forest due to fertilisation. General improvement of living conditions by narrowing the C/N ratio. [85
Landfill soil high NH4Cl and KNO3 Lower methane oxidation when N becomes limiting; addition of ammonium and nitrate stimulated methane consumption. N limitation. [86
Landfill soil high NH4Cl, KNO3, lime Rapid onset of methane oxidation with ammonium and nitrate in soil columns; nitrate also stimulated in fresh landfill soil. None. [36
Coniferous forest soil low CaNH4NO3 Higher methane oxidation in forest soils which have been previously fertilised (2 years). None. [87
Rice field soil high NH4Cl Inhibition turns into stimulation by ammonium after several incubations with different methane mixing ratios. Nitrifiers are stimulated and are therefore consuming more methane. [43
Landfill cover soil high NH4Cl, nitrified sludge, compost Stimulation of methane oxidation by ammonium of soil exposed to high methane for short period; inhibition increases with exposure time to high methane. Change in the community composition to N-susceptible methanotrophs. [37
Tropical pasture soil low (NH4)2SO4, urea, CaNO3 Fertilised pastures had higher methane uptake than traditional and legume pastures. Higher plant biomass in fertilised pastures results in lower water filled pore space and hence higher diffusion of methane into the soil. [35
Rice field soil high urea Stimulation of in vitro methane oxidation. N-limitation of MOB. [77
Natural (forest/prairie) and agricultural soils low various regimes Positive correlation between mineral N of soil and methane consumption at elevated levels. Increasing contribution of nitrifiers to methane oxidation. [88
N-limited forest soil low (NH4)2SO4 Higher methane uptake in N-limited soils after initial short-term inhibition. N limitation of the methanotrophs. [20
Landfill cover soil high NH4Cl, (NH4)2SO4 Stimulation of methane oxidation at high methane (2%) only when sufficient N was present. Nitrogen limitation of type I methanotrophs. [38
Agricultural soil low various regimes High levels of fertilisation (150 kg N ha−1 year−1) resulted in higher methane uptake of cultivated soils. Higher growth of methanotrophs; fertilised plots had not been tilled for a long period leading to different soil pore structure and better methane diffusion into the soil. [89

Presented are the habitats, whether these habitats are characterised by high or low methane concentrations, which fertiliser was used if any, what the nature of the observed effects was and which explanation the authors gave for the latter.

The explanations presented in these studies were rather diverse (see Table 2). Some authors did not suggest an explanation, or failed to discuss the findings at all [36,78,83,84]. Others claimed that increased methane consumption by ammonia oxidisers [43,88], changes in community composition of the methanotrophic bacteria [37], general improvement of nutrient availability or methane diffusivity through enhanced development of the vegetation may have been responsible for the observed effects [35,79,85]. Surprisingly, only Bender and Conrad [81], De Visscher et al. [86] and Papen and co-workers [20] argued that methane oxidisers simply need an N source and that the observed effects were a relief of a limitation rather than a stimulation of activity. Already in 1995, Bender and Conrad noted that their results were rather surprising given the fact that many studies reported inhibition only. However, the surprising fact is not that methanotrophic bacteria need nitrogen, but that the observations (summarised in Table 2) have never been followed up experimentally in order to assess environmental consequences of reduced methane consumption under N-limiting conditions. The studies were performed only to investigate the dimensions of the detrimental effect of nitrogen fertilisation on the methane sink function of soils. Yet, one would think that the preservation or even increase of the sink strength due to the relief of a nutrient limitation would be equally important to study in more detail. Very recently, De Visscher and co-workers [38] published the first study that anticipated both inhibitory and stimulatory effects of ammonium on methane oxidation. Nevertheless, experimental proof for a mechanism of nitrogen-based stimulation of methane oxidation in soil is still missing.

Nitrogen-based stimulation of methane oxidation in soil: hypothetical mechanisms

Since none of the studies discussed in the preceding paragraph presents conclusive experimental evidence for the mechanisms underlying the stimulatory effect of nitrogen on methane oxidation, some speculation is necessary. Basically, the addition of nitrogenous fertiliser can act directly on cellular level or it may evoke changes in the soil ecosystem that influence methanotrophic bacteria indirectly. There are three options with respect to the former: (1) the ammonium or nitrate relieves N limitation of cell growth and subsequently increases the activity of methanotrophic community on the long term; (2) N addition interferes more directly with the synthesis of involved enzymes in the methane oxidation pathway of nitrogen-starved cells; (3) the size and activity of the nitrifying population, which also oxidises methane, is increased.

Methanotrophic bacteria have a relatively high nitrogen requirement. For every mole of assimilated carbon 0.25 moles of N have to be taken up [90]. The assimilatory demand for ammonium by methanotrophic bacteria has been demonstrated to even suppress ammonia oxidation in soils [91]. Hence, especially in environments where the molar ratio of methane to nitrogen is higher than 10 (assuming 40% assimilation of every mole of consumed methane), such as in the rhizosphere of wetland plants, in landfill soils and in the upper layer of non-eutrophic sediments, N limitation may occur. Long-term depletion of an N source will inevitably reduce protein synthesis and growth, leading to a reduction or even cessation of methane consumption. Potentially, this limitation can be compensated for by the fixation of molecular nitrogen in these habitats. Type II and type X methanotrophic bacteria have been demonstrated to fix molecular nitrogen [92], while recently type I methanotrophic bacteria have also been shown to contain genes coding for nitrogen fixation pathway [93]. Hence, the absence of ammonium or nitrate per se does not necessarily imply reduced methane oxidation. However, nitrogen fixation is a costly process in terms of energy and reducing equivalents, and switching from ammonium or nitrate uptake to nitrogen fixation will result in lower growth potential of the methanotrophic bacteria, and possibly lower methane oxidation rates. Additions of ammonium or nitrate to nitrogen-fixing methanotrophic communities in soil could thus lead to stimulated methane oxidation due to the switch to the energetically more favourable consumption of inorganic nitrogen.

Experimental evidence from rice soils [9] but also from culture experiments [94] strongly suggests that a more direct mechanism of stimulation of methane oxidation by inorganic nitrogen compounds might exist than just the relief of nitrogen limitation for growth.

Methane oxidisers from the nitrogen-depleted rhizosphere of rice display a lag in activity of more than 2 days [9]. However, the addition of ammonium leads to immediate methane consumption in methane oxidation assays. Moreover, methane emission from these rice microcosms displayed a very strong inverse relationship with the ammonium availability in the rhizosphere of the plants (Fig. 2), indicating a continuous response of methane consumption to ammonium. Park et al. [94] observed immediate loss of particulate MMO activity upon nitrate depletion in a batch culture of Methylosinus trichosporium, while activity was restored again within hours following nitrate application. These observations cannot be explained by stimulated growth but only by a more direct effect of ammonium or nitrate on the methane-consuming metabolism itself. Since the immediate stimulation of methane oxidation is mediated by both ammonium and nitrate, a mechanistic explanation must be linked to nitrogen assimilation. The question is how the assimilation of nitrogen is connected so rapidly to the dissimilation of methane. We propose the following mechanism that involves nitrogen fixation. In Fig. 3a the methane oxidation pathway is schematically presented in the case of non-nitrogen-limiting conditions. The first step in this pathway requires reducing equivalents in the form of NADH2. This reduced compound is derived mainly from the oxidation of formate by formate dehydrogenase [95]. However, the NADH2 produced in this step has also been demonstrated to serve as electron donor for methanotrophic nitrogenase activity (cf. [95]). The electron flow is mediated by ferredoxin (FD)-NAD+ oxidoreductase and FD. This diversion of electron flow towards FD is blocked when ammonium or nitrate is assimilated under nitrogen excess conditions. Switching between inorganic nitrogen assimilation and nitrogen fixation in bacteria can proceed rapidly and is under control of ntr gene control systems [96], which regulates nitrogen assimilation and nitrogen fixation at the transcriptional level. In Methylococcus capsulatus, nitrogen fixation was switched off 5 min after addition of the ammonia [97]. Glutamine was the possible internal regulatory molecule, indicating that an ntr type of control was involved [96]. When no ammonium or nitrate is available (Fig. 3b) the ntr system will allow the nitrogenase system to be active, thereby diverting NADH2 to nitrogen fixation. Combined with the fact that methylotrophs have a high NADH2 requirement for C assimilation [98], this could lead to a limited supply of NADH2 to the monooxygenase. Consequently, the oxidation of methane proceeds at a level that is often below the detection limit of the common methane consumption assays. This hypothetical mechanism would explain the rapid responses of methane oxidation to ammonium and nitrate addition that are observed in cultures, soil incubations and field studies. This mechanism is most likely to occur when methane consumption and hence NADH2 generation are low, and there is a limited availability of nitrogen. Also, elevated C assimilation relative to respiration could lead to a limited supply of NADH2 to the monooxygenase enzyme [88,98]. The rhizosphere of wetland plants or nitrogen-limited upland and landfill soils are the environments where this mechanism could operate. However, the proposed mechanism is only one hypothetical possibility. Of course other modes of connection between the nitrogen assimilation pathways and the monooxygenase enzyme machinery, including gene transcriptional regulation, have to be considered.

2

Relationship between the pore water ammonium concentration in the rhizosphere soil of rice soil microcosms and methane emission from these systems. These microcosms were supplemented weekly after pore water samples had withdrawn with (NH4)2PO4 to a total amount of 400 kg N ha−1. The emission measurements were analysed weekly for a period of 10 weeks. The dotted line indicates 95% confidence interval of the fitted linear regression.

2

Relationship between the pore water ammonium concentration in the rhizosphere soil of rice soil microcosms and methane emission from these systems. These microcosms were supplemented weekly after pore water samples had withdrawn with (NH4)2PO4 to a total amount of 400 kg N ha−1. The emission measurements were analysed weekly for a period of 10 weeks. The dotted line indicates 95% confidence interval of the fitted linear regression.

3

Schematic presentation of the hypothetical mechanism explaining the immediate stimulation of methane oxidation by ammonium or nitrate addition to soils or sediments. a: The flow of NADH2 when nitrogen is not limiting. NADH2 is diverted to the MMO reaction. The ntr gene control system prevents the use of NADH2 for the nitrogenase reaction. b: The situation when inorganic nitrogen is available for assimilation. The ntr gene control system diverts NADH2 to enable nitrogen fixation thereby reducing or even diminishing the NADH2 flow to the methane oxidation reaction.

3

Schematic presentation of the hypothetical mechanism explaining the immediate stimulation of methane oxidation by ammonium or nitrate addition to soils or sediments. a: The flow of NADH2 when nitrogen is not limiting. NADH2 is diverted to the MMO reaction. The ntr gene control system prevents the use of NADH2 for the nitrogenase reaction. b: The situation when inorganic nitrogen is available for assimilation. The ntr gene control system diverts NADH2 to enable nitrogen fixation thereby reducing or even diminishing the NADH2 flow to the methane oxidation reaction.

Since ammonia-oxidising bacteria also have the potential to consume methane, the stimulatory effect of nitrogen additions could also be related to enhanced populations of nitrifiers, and subsequent enhanced methane oxidation by these organisms. However, the phenomena as observed in the rice microcosm experiments are most definitely not related to ammonia oxidisers. First of all the numbers of methanotrophic bacteria increased due to the fertilisation (see also Fig. 1), indicating that they have been utilising methane while the ammonia-oxidising bacteria did not increase [9]. Secondly, using data from culture experiments [99], it would have required between 108 and 109 cells per g of dry soil to account for the observed methane oxidation rate. Most probable number counts in these rice soils yielded between 105 and 106 ammonia oxidisers per g [9]. Moreover, the stimulatory effects of inorganic nitrogen additions in high methane environments like landfill soils were also demonstrated to occur with nitrate [86], ruling out an involvement of ammonia oxidisers. This also appears to be the case for atmospheric methane uptake. The observed positive relationship between ammonium content of soils and methane consumption (Table 2) would require an unrealistically high number of cells, if it were caused by ammonia oxidisers [100].

The overview in Table 2 offers some explanations of nitrogen-based stimulation of methane oxidation that acts indirectly through changes in the habitat of the methanotrophic bacteria. Fertilisation leads to enhanced development of the vegetation. The subsequent increased evapotranspiration lowers the soil water-filled pore space, leading to higher diffusion of methane and oxygen into the soil [35,89]. Of course, the studies proposing this explanation have been performed in upland soils. In wetlands this would be a highly unlikely mechanism. It has been suggested that enhanced biomass of wetland vegetation can lead to higher soil oxygen input from the roots of these plants, which combined with higher nitrogen availability stimulates methane oxidation [101,102].

Fertiliser-induced effects on the vegetation can also lead to a general improvement of soil fertility, creating improved conditions for microbial growth. Higher mineralisation rates will lead to higher carbon, nitrogen and possibly micronutrient availability [79,80,85]. The improved availability of carbon may be especially important for atmospheric methane consumers, which have been demonstrated to profit from and may even depend on additional carbon sources besides methane [103,104]. Changes in composition of the vegetation can also lead to altered soil texture around the new root systems, which may result in improved diffusivity of gases into the soil.

It is evident that nitrogenous fertiliser additions may promote methane uptake in both upland and lowland soils directly or indirectly. However, direct experimental evidence substantiating or ruling out any of the mechanisms described above is missing. It will be challenging to unravel the relationships between nitrogen availability and methane consumption, and to identify involved bacteria. For the time being nitrogen has to be treated as a potentially inhibitory and as a beneficial factor for methane consumption in soils and sediments.

Environmental aspects of the N prerequisite of methane consumption in soils and sediments

The reduced methane consumption activity under N-limiting conditions can have major environmental consequences. The methane source strength of soils and sediments can be enhanced while the sinks of atmospheric methane may loose part of their methane mitigating potential. However, this will depend on the natural nitrogen dynamics and also on the anthropogenic nitrogen input into soils and sediments through fertilisers and industry-derived atmospheric deposition.

Nitrogen regulation of methane emission from rice paddies and natural wetlands

Rice paddies

Rice paddies are among the most prominent methane sources on earth [3]. The contribution of rice paddies is believed to increase in future as the consequence of increased use of nitrogenous fertiliser in order to increase crop yield. More fertiliser will inhibit methane oxidation and enhance emission; at least that was the general idea. However, experimental evidence assessing this issue has been contradictory. Lower as well as higher methane emissions have been found following fertiliser application (cf. [9]). The unexpected reduction of methane emission after fertiliser application has often been explained by inhibition of methanogenesis. However, the observations of reduced emission may also be explained by a stimulation of methane consumption due to the elimination of the N-limiting conditions for the methanotrophic bacteria. Overlooking this possibility may be the result of the applied methodology for assessing methane oxidation in these systems. Often inhibitors of methane oxidation were used to estimate the percentage of methane oxidised in the rhizosphere of rice plants that also inhibited methanogenesis and led to overestimation of methane consumption. As already outlined in Section 3, Krüger and co-workers [10,76] employed for the first time a ‘truly’ selective inhibitor (difluoromethane; CH2F2) of methane oxidation, and therefore their measurements were the first reliable estimates of the actual methane oxidation in natural rice fields. These studies indicated that increased use of fertiliser would lead to lowering of the methane emission, a fact that has to be considered in global methane emission models. A very recent study using 13C-labelled CH4 in rice microcosms confirmed this [105].

Variability of methane fluxes in relation to environmental dynamics or agricultural practices (e.g. fertilisation) have generally been assessed top-down in an ecosystem fashion, with methane emission or consumption as the response variable. The underlying microbial processes, in particular the characteristics and ecology of involved microorganisms, have seldom been taken into account when explaining observed variability in global methane budgets. The microcosm experiments performed by Bodelier and co-workers [8] even indicated that the diversity of the methanotrophic bacteria might play a major role in this issue. Genera belonging to the type I methanotrophic bacteria were preferentially stimulated by the addition of N, indicating that the characteristics of involved organisms have to be taken into account when explaining global fluxes. Therefore, to better understand regulation of methane oxidation by nitrogen in rice paddies, in situ oxidation technique of Krüger and co-workers should be generally applied, and combined with the physicochemical characterisation of the soil and community analyses of methane oxidisers.

Natural wetlands

Similarly as in rice paddies, the dimension and dynamics of methane emission in natural wetlands (e.g. bogs, fens, swamps, marshes) are of high environmental importance [3]. The consequences of anthropogenic disturbances receive a lot of attention. Rhizospheric methane oxidation in natural wetlands has been demonstrated to fluctuate during the growing season of wetland plants and also among plant species [101,106]. High oxidation rates were detected early in the plant growth cycle and low rates when plants matured. The gradually diminishing nitrogen pool for methane oxidation, caused by increased nitrogen uptake by the growing plants, may well explain these observations. Likewise, differences in rhizospheric methane oxidation among plant species may be explained by different nitrogen demands of the plants. The decrease of ammonium in pore water during the growing season has been documented for rice paddies [9,76], and there is no reason to assume that this will be different for natural wetlands that are normally unfertilised.

The distinctly enhanced deposition of atmospheric nitrogen over the past decades, which is still increasing, can be regarded as nitrogen fertilisation. Especially in nutrient-poor mires and peat lands this could result in a shift to more productive plant species, leading to higher methane production and emission from these systems. However, Granberg and co-workers [102] found a strong negative effect of nitrogen (NH4NO3) additions on methane emission from boreal mire. This effect was found only when the cover of plants (Carex sp.) was high. The high plant density probably provides sufficient oxygen to the methanotrophic bacteria to profit from the nitrogen additions. In northern peat lands, Updegraff [107] found a strong correlation between methane flux and nitrogen retention. Methane emissions from fens were much lower than from bogs, a trend that the authors attributed to the higher nitrogen availability in the fens due to the lower plant productivity. The authors concluded that stimulation of methane oxidation by nitrogen was the most likely explanation for the differences in methane emissions between bogs and fens.

Hence, the understanding of the dynamics of methane fluxes from rice paddies and natural wetlands, especially in response to anthropogenic disturbances, has been misguided by the strong focus on nitrogen inhibition aspects on the methane consumption side and by the inability to assess the in situ activity of methanotrophic bacteria. Future studies assessing the effect of more intensive fertiliser use, enhanced atmospheric N deposition, elevated temperatures and CO2 on methane emission from wetlands should also assess these questions bottom-up determining the in situ functioning and diversity of methanotrophic bacteria. The latter will include the combined use of specific inhibitors together with the techniques that make use of incorporation of stable isotopes (13CH4) into phylogenetically relevant markers (phospholipid-derived fatty acid (PLFA), DNA, RNA) [108–110].

Nitrogen regulation of sink strength of upland soils

The consumption of atmospheric methane by upland soils can be affected by a vast array of environmental factors, as has been extensively investigated and reviewed [3,7,17]. Atmospheric nitrogen deposition in forest soils has received a lot of attention and is still a matter of great concern. As seen in Table 2, there are several studies on forest soils that contradict the generally found suppressing effect. Hence, even in the case of forest soils, nitrogen has to be taken into account as a possible stimulus for methane oxidation. It is hard to imagine, however, that atmospheric methanotrophic bacteria are actually limited by N. Methane concentrations in upland soils are in the nM range and up to date it is still not clear which bacteria are responsible for the process and whether growth on atmospheric methane alone is possible (cf. [111]). It has been demonstrated though that the consumption of atmospheric methane in soils can be promoted by non-methane substrates like methanol, formate and acetate [102,103,111]. If methane is not the primary source of carbon and energy for these microorganisms, then nitrogen input in upland soils may facilitate mixotrophic or even heterotrophic growth of bacteria capable of atmospheric methane consumption. The resulting elevated cell numbers will cause higher consumption of atmospheric methane. This would be a possible explanation for the positive correlation between ammonium content and atmospheric methane consumption in upland soils. Another explanation for the latter involves the assumption that the atmosphere is the only source of methane for high-affinity methanotrophic bacteria, a matter that is still not resolved. All upland soils may become partially anoxic following high precipitation events and start producing methane. The methanotrophic bacteria can profit from this enhanced methane flux and grow when sufficient nitrogen is present. Thereby the nitrogen status of upland soils may allow for a population increase of microorganisms, which, after drying of the soil, retain a higher potential for atmospheric methane consumption.

Hence, the sink strength of upland soils for atmospheric methane can be regulated by nitrogen. However, whether this effect is negative or positive, has to be assessed on a case-wise basis. There are numerous physical, chemical and management aspects that make comparisons between soils in light of one single factor (fertiliser or N deposition) difficult to interpret. It is clear, however, that nitrogen is a potential regulating factor that has to be investigated in more detail if we are to properly assess the role of upland soils in the global methane budget.

Nitrogen regulation of the biofilter function of soils

Landfills are prominent sources of methane, emitting 10–70 Tg of CH4 per year (cf. [86]) globally. Covering the landfills with an aerobic soil layer containing methanotrophic bacteria reduces methane emission. The effect of nitrogenous additions to these cover soils has also been the subject of numerous studies, again bearing in mind the possible reduction of methane consumption and subsequent enhanced emission. In recent years it has been shown that in landfill cover soils the addition of ammonium or nitrate can stimulate methane oxidation [36,37,86]. The rapid stimulation found in these studies points to an enzymatic effect rather than a growth-related response. This could be mediated through the proposed mechanisms listed in Fig. 3. Nevertheless, the management of landfill cover soils in order to reduce methane emission as effectively as possible has to take into account the nitrogen requirement of methanotrophic bacteria. However, landfill cover soils are normally vegetated, which can result in nitrogen-limiting conditions and hence, higher methane emissions. Bare landfill cover soils are not favoured since the plants are required for the prevention of erosion of the cover soil. Therefore, fertilisation strategies need to be developed in order to ensure optimal methane oxidation.

The top soil layer of C-rich wet soils can also act as a biofilter for methane. Kammann and co-workers [69] demonstrated a high methane production potential for managed grassland after anoxic incubation, which may occur after wetting of the soil. The methane concentration in these soils increased substantially after autumn rainfall. Nevertheless, no methane was released due to the very active methane-oxidising community in the top soil layer. Nitrogen limitation of methane oxidation in the top layers of wet soils may turn these systems into a source of methane and therefore also these soil systems are an object for future investigations on the regulatory role of N in methane consumption.

Conclusion

The important role of methanotrophic bacteria in the global methane budget, and hence in our present and future climate, is evident. This realisation has resulted in an immense amount of studies on the reactions and fluxes that these bacteria catalyse, as well as on the factors that control, regulate and affect this process. Nitrogenous fertilisers have generally been regarded as inhibitory to methane consumption by soils and sediments. This paradigm, combined with traditional ‘top-down’ ecosystem approach used in global methane flux studies, has led to erroneous interpretation of data by ignoring the ecological characteristics of involved organisms. With respect to methanotrophic bacteria the knowledge is still far from complete. The essential role of mineral nitrogen availability for these microorganisms and the process they mediate, have been largely neglected. Mineral nitrogen seems to be a prerequisite for the occurrence of methane consumption and might even initiate and stimulate the enzymatic machinery in a yet unknown way. The stimulation differentially affects methanotrophic species, which demonstrates the essence of a ‘bottom-up’ approach in global flux studies. These facts place methane oxidation in soils and sediments into a new perspective. New research approaches are required to link nitrogen availability with methanotrophic bacteria. Soil physicochemical and biological factors (e.g. competition for N with plants and other microorganisms) are potential areas for new research. However, we first have to define new concepts and research questions that will have to integrate ‘top-down’ ecosystem studies with ‘bottom-up’ approaches considering microbial populations and cellular characteristics.

Acknowledgements

The authors would like to thank Dr. Jay Gulledge and an anonymous referee for the reviewing of this manuscript. P.L.E.B. was financially supported by a grant of the Centre for Wetland Ecology (http://www.kun.nl/waterkracht/cwe). This paper is publication no. 3216 of the NIOO-KNAW, Centre for Limnology, and publication no. 349 of the Centre for Wetland Ecology (CWE).

Abbreviations

    Abbreviations
  • MMO

    methane monooxygenase

  • FD

    ferredoxin

References

[1]
IPCC
(
2001
)
Climate change 2001: The scientific basis
.
Contribution of working group I to the third assessment report of the Intergovernmental panel on climate change
  (
Houghton
J.T.
Ding
Y.
Griggs
D.J.
Noguer
M.
Vanderlinden
P.J.
Dai
X.
Maskell
K.
Johnson
C.A.
, Eds.), p.
881
.
Cambridge University Press
,
Cambridge
.
[2]
Crutzen
P.J.
(
1995
)
The role of methane in atmospheric chemistry and climate
. In:
Ruminant Physiology: Digestion, Metabolism, Growth and Reproduction: Proceedings of the Eighth International Symposium on Ruminant Physiology
  (
Engelhardt
W.V.
Leonhard-Marek
S.
Breves
S.
Giesecke
D.
, Eds.), pp.
291
315
.
Ferdinand Enke Verlag
,
Stuttgart
.
[3]
Le Mer
J.
Roger
P.
(
2001
)
Production, oxidation, emission and consumption of methane by soils: a review
.
Eur. J. Soil Biol.
 
37
,
25
50
.
[4]
Hanson
R.S.
Hanson
T.E.
(
1996
)
Methanotrophic bacteria
.
Microbiol. Rev.
 
60
,
439
471
.
[5]
Frenzel
P.
(
2000
)
Plant-associated methane oxidation in rice fields and wetlands
. In:
Advances in Microbial Ecology
  (
Schink
B.
, Ed.), pp.
85
114
.
Kluwer Academic/Plenum Publishers
,
New York
.
[6]
Steudler
P.A.
Bowden
R.D.
Melillo
J.M.
Aber
J.D.
(
1989
)
Influence of nitrogen fertilisation on methane uptake in forest soils
.
Nature
 
341
,
314
316
.
[7]
Hütsch
B.W.
(
2001
)
Methane oxidation in non-flooded soils as affected by crop production – invited paper
.
Eur. J. Agron.
 
14
,
237
260
.
[8]
Bodelier
P.L.E.
Roslev
P.
Henckel
T.
Frenzel
P.
(
2000
)
Stimulation by ammonium-based fertilisers of methane oxidation in soil around rice roots
.
Nature
 
403
,
421
424
.
[9]
Bodelier
P.L.E.
Hahn
A.P.
Arth
I.R.
Frenzel
P.
(
2000
)
Effects of ammonium-based fertilisation on microbial processes involved in methane emission from soils planted with rice
.
Biogeochemistry
 
51
,
225
257
.
[10]
Krüger
M.
Frenzel
P.
(
2003
)
Effects of N-fertilization on CH4 oxidation and production, and consequences for CH4 emissions from microcosms and rice fields
.
Glob. Change Biol.
 
9
,
773
784
.
[11]
Bender
M.
Conrad
R.
(
1992
)
Kinetics of CH4 oxidation in oxic soils exposed to ambient air or high mixing ratios
.
FEMS Microbiol. Ecol.
 
101
,
261
270
.
[12]
Dunfield
P.F.
Conrad
R.
(
2000
)
Starvation alters the apparent half-saturation constant for methane in the type II methanotroph Methylocystis strain LR1
.
Appl. Environ. Microbiol.
 
66
,
4136
4138
.
[13]
Henckel
T.
Jäckel
U.
Schnell
S.
Conrad
R.
(
2000
)
Molecular analyses of novel methanotrophic communities in forest soil that oxidize atmospheric methane
.
Appl. Environ. Microbiol.
 
66
,
1801
1808
.
[14]
Holmes
A.J.
Roslev
P.
MacDonald
I.R.
Iversen
N.
Henriksen
K.
Murrell
J.C.
(
1999
)
Characterisation of methanotrophic bacterial populations in soils showing atmospheric methane uptake
.
Appl. Environ. Microbiol.
 
65
,
3312
3318
.
[15]
Bull
I.D.
Parekh
N.P.
Hall
G.H.
Ineson
P.
Evershed
R.P.
(
2000
)
Detection and classification of atmospheric methane oxidising bacteria in soil
.
Nature
 
405
,
175
178
.
[16]
Gulledge
J.
Schimel
J.P.
(
1998
)
Low-concentration kinetics of atmospheric CH4 oxidation in soil and mechanism of NH4+ inhibition
.
Appl. Environ. Microbiol.
 
64
,
4291
4298
.
[17]
Smith
K.A.
Dobbie
K.E.
Ball
B.C.
Bakken
L.R.
Sitaula
B.K.
Hansen
S.
Brumme
R.
Borken
W.
Christensen
S.
Priemé
A.
Fowler
D.
MacDonald
J.A.
Skiba
U.
Klemedtson
L.
Kasimir-Klemedtsson
A.
Degórska
A.
Orlanski
P.
(
2000
)
Oxidation of atmospheric methane in Northern European soils, comparison with other ecosystems, and uncertainties in the global terrestrial sink
.
Glob. Change Biol.
 
6
,
791
803
.
[18]
King
G.M.
Schnell
S.
(
1994
)
Effect of increasing atmospheric methane concentration on ammonium inhibition of soil methane consumption
.
Nature
 
370
,
282
284
.
[19]
Steinkamp
R.
Butterbach-Bahl
K.
Papen
H.
(
2001
)
Methane oxidation by soils of an N-limited and N-fertilized spruce forest in the Black Forest, Germany
.
Soil Biol. Biochem.
 
33
,
145
153
.
[20]
Papen
H.
Daum
M.
Steinkamp
R.
Butterbach-Bahl
K.
(
2001
)
N2O and CH4-fluxes from soils of a N-limited and N-fertilised spruce forest ecosystem of the temperate zone
.
J. Appl. Bot.
 
75
,
159
163
.
[21]
Nesbit
S.P.
Breitenbeck
G.A.
(
1992
)
A laboratory study of factors influencing methane uptake by soils
.
Agricult. Ecosyst. Environ.
 
41
,
39
54
.
[22]
Whalen
S.C.
(
2000
)
Influence of N and non-N salts on atmospheric methane oxidation by upland boreal forest and tundra soils
.
Biol. Fertil. Soils
 
31
,
279
287
.
[23]
Philips
R.L.
Whalen
S.C.
Schlesinger
W.H.
(
2001
)
Response of soil methanotrophic activity to carbon dioxide enrichment in a North Carolina coniferous forest
.
Soil Biol. Biochem.
 
33
,
793
800
.
[24]
Bradford
M.A.
Ineson
P.
Wookey
P.A.
Lappin-Scott
H.M.
(
2001
)
the effects of acid nitrogen and acid sulphur deposition on CH4 oxidation in a forest soil: a laboratory study
.
Soil Biol. Biochem.
 
33
,
1695
1702
.
[25]
Wang
Z.P.
Ineson
P.
(
2003
)
Methane oxidation in a temperate coniferous forest soil: effects of inorganic N
.
Soil Biol. Biochem.
 
35
,
427
433
.
[26]
Adamsen
A.P.S.
King
G.M.
(
1993
)
Methane consumption in temperate and subarctic forest soils: rates, vertical zonation, and responses to water and nitrogen
.
Appl. Environ. Microbiol.
 
59
,
485
490
.
[27]
Bronson
K.F.
Mosier
A.R.
(
1994
)
Suppression of methane oxidation in aerobic soil by nitrogen fertilisers, nitrification inhibitors, and urease inhibitors
.
Biol. Fertil. Soil
 
17
,
263
268
.
[28]
Hütsch
B.W.
Russell
P.
Mengel
K.
(
1996
)
CH4 oxidation in two temperate arable soils as affected by nitrate and ammonium application
.
Biol. Fertil. Soils
 
23
,
86
92
.
[29]
Hütsch
B.W.
(
1998
)
Methane oxidation in arable soil as affected by ammonium, nitrite, and organic manure with respect to soil pH
.
Biol. Fertil. Soils
 
28
,
27
35
.
[30]
Kravchenko
I.
Boeckx
P.
Galchenko
V.
Van Cleemput
O.
(
2002
)
Short- and medium-term effects of NH4+ on CH4 and N2O fluxes in arable soils with different texture
.
Soil Biol. Biochem.
 
34
,
669
678
.
[31]
Sitaula
B.K.
Hansen
S.
Sitaula
J.I.B.
Bakken
L.R.
(
2000
)
Methane oxidation potentials and fluxes in agricultural soil: effects of fertilisation and soil compaction
.
Biogeochemistry
 
48
,
323
339
.
[32]
Hansen
S.
Maehlum
J.E.
Bakken
L.R.
(
1993
)
N2O and CH4 fluxes in soil influenced by fertilization and tractor traffic
.
Soil Biol. Biochem.
 
25
,
621
630
.
[33]
Dunfield
P.F.
Knowles
R.
(
1995
)
Kinetics of inhibition of methane oxidation by nitrate, nitrite, and ammonium in a humisol
.
Appl. Environ. Microbiol.
 
61
,
3129
3135
.
[34]
Cril
P.M.
Martikainen
P.J.
Nykanen
H.
Silvola
J.
(
1994
)
Temperature and N fertilization effects on methane oxidation in a drained peatland soil
.
Soil Biol. Biochem.
 
26
,
1331
1339
.
[35]
Veldkamp
E.
Weitz
A.M.
Keller
M.
(
2001
)
Management effects on methane fluxes in humid tropical pasture soils
.
Soil Biol. Biochem.
 
33
,
1493
1499
.
[36]
Hilger
H.A.
Wollum
A.G.
Barlaz
M.A.
(
2000
)
Landfill methane oxidation response to vegetation, fertilisation, and liming
.
J. Environ. Qual.
 
29
,
324
334
.
[37]
de Visscher
A.
Schippers
M.
VanCleemput
O.
(
2001
)
Short-term response of enhanced methane oxidation in landfill cover soils to environmental factors
.
Biol. Fertil. Soils
 
33
,
231
237
.
[38]
de Visscher
A.
Van Cleemput
O.
(
2003
)
Induction of enhanced CH4 oxidation in soils: NH4+ inhibition patterns
.
Soil Biol. Biochem.
 
35
,
907
913
.
[39]
Kightley
D.
Nedwell
D.B.
Cooper
M.
(
1995
)
Capacity for methane oxidation in landfill cover soils measured in laboratory-scale soil microcosms
.
Appl. Environ. Microbiol.
 
61
,
592
601
.
[40]
van der Nat
F.J.W.A.
de Brouwer
J.F.C.
Middelburg
J.J.
Laanbroek
H.J.
(
1997
)
Spatial distribution and inhibition by ammonium of methane oxidation in intertidal freshwater marshes
.
Appl. Environ. Microbiol.
 
63
,
4734
4740
.
[41]
Conrad
R.
Rothfuss
F.
(
1991
)
Methane oxidation in the soil surface layer of a flooded rice field and the effect of ammonium
.
Biol. Fertil. Soils
 
12
,
28
32
.
[42]
Cai
Z.
Yan
X.
(
1999
)
Kinetic model for methane oxidation by paddy soil as affected by temperature, moisture and N addition
.
Soil Biol. Biochem.
 
31
,
715
725
.
[43]
Cai
Z.C.
Mosier
A.R.
(
2000
)
Effect of NH4Cl addition on methane oxidation by paddy soils
.
Soil Biol. Biochem.
 
32
,
1537
1545
.
[44]
Cai
Z.
Mosier
A.R.
(
2002
)
Restoration of methane oxidation abilities of desiccated paddy soils after re-watering
.
Biol. Fertil. Soils
 
36
,
183
189
.
[45]
Bosse
U.
Frenzel
P.
Conrad
R.
(
1993
)
Inhibition of methane oxidation by ammonium in the surface layer of a littoral sediment
.
FEMS Microbiol. Ecol.
 
13
,
123
134
.
[46]
King
G.M.
(
1990
)
Dynamics and controls of methane oxidation in a Danish wetland sediment
.
FEMS Microbiol. Ecol.
 
74
,
309
324
.
[47]
Hütsch
B.W.
Webster
C.P.
Powlson
D.S.
(
1993
)
Long-term effects of nitrogen-fertilisation on methane oxidation in soil of the Broadbalk wheat experiment
.
Soil Biol. Biochem.
 
26
,
1307
1315
.
[48]
Gulledge
J.
Doyle
A.P.
Schimel
J.P.
(
1997
)
Different NH4+-inhibition patterns of soil CH4-oxidisers populations across sites
Soil Biol. Biochem.
 
29
,
13
21
.
[49]
Butterbach-Bahl
K.
Gasche
R.
Huber
C.H.
Kreutzer
K.
(
1998
)
Impact of N-input by wet deposition on N-trace gas fluxes and CH4-oxidation in spruce forest ecosystems of the temperate zone in Europe
.
Atmos. Environ.
 
32
,
559
564
.
[50]
Gulledge
J.
Schimel
J.P.
(
2000
)
Controls on soil carbon dioxide and methane fluxes in a variety of Taiga forest stands in interior Alaska
.
Ecosystems
 
3
,
269
282
.
[51]
Butterbach-Bahl
K.
Breuer
L.
Gasche
R.
Willibald
G.
Papen
H.
(
2002
)
Exchange of trace gases between soils and the atmosphere in Scots pine forest ecosystems of the north eastern German lowlands 1: Fluxes of N2O, NO/NO2 and CH4 at forest sites with different N-deposition
.
Forest Ecol. Manage.
 
123
,
123
134
.
[52]
MacDonald
J.A.
Skiba
U.
Sheppard
L.J.
Hargreaves
K.J.
Smith
K.A.
Fowler
D.
(
1996
)
Soil environmental variables affecting the flux of methane from a range of forest, moorland and agricultural soils
.
Biogeochemistry
 
34
,
113
132
.
[53]
Castro
M.S.
Steudler
P.A.
Melillo
J.M.
Aber
J.D.
Millham
S.
(
1993
)
Exchange of N2O and CH4 between the atmosphere and soils in spruce-fir forests in the northeastern United States
.
Biogeochemistry
 
18
,
119
135
.
[54]
Saari
A.
Martikainen
P.J.
Ferm
A.
Ruuskanen
J.
de Boer
W.
Troelstra
S.R.
Laanbroek
H.J.
(
1997
)
Methane oxidation in soil profiles of Dutch and Finnish coniferous forests with different texture and atmospheric nitrogen deposition
.
Soil Biol. Biochem.
 
29
,
1625
1632
.
[55]
Hütsch
B.W.
(
1996
)
Methane oxidation in soils of two long-term fertilisation experiments in Germany
.
Soil Biol. Biochem.
 
28
,
773
782
.
[56]
Mosier
A.R.
Schimel
D.
Valentine
D.
Bronson
K.
Parton
W.
(
1991
)
Methane and nitrous oxide fluxes in native, fertilised and cultivated grasslands
.
Nature
 
350
,
330
332
.
[57]
Willison
T.W.
Webster
C.P.
Goulding
K.W.T.
Powlson
D.S.
(
1995
)
Methane oxidation in temperate soils: effects of land use and the chemical form of nitrogen fertiliser
.
Chemosphere
 
30
,
539
546
.
[58]
Neff
J.C.
Bowman
W.D.
Holland
E.A.
Fisk
M.C.
Schmidt
S.K.
(
1994
)
Fluxes of nitrous oxide and methane from nitrogen-amended soils in a Colorado alpine ecosystem
.
Biogeochemistry
 
27
,
23
33
.
[59]
Weitz
A.M.
Keller
M.
(
1999
)
Spatial and temporal variability of nitrogen and methane fluxes from a fertilised tree plantation in Costa Rica
.
J. Geophys. Res.
 
104
,
30097
30107
.
[60]
Whalen
S.C.
Reeburgh
W.S.
(
2000
)
Effect of nitrogen fertilisation on atmospheric methane oxidation in boreal forest soils
.
Chemosphere
 
2
,
151
155
.
[61]
Bradford
M.A.
Wookey
P.A.
Ineson
P.
Lappin-Scott
H.M.
(
2001
)
Controlling factors and effects of chronic nitrogen and sulphur deposition on methane oxidation in a temperate forest soil
.
Soil Biol. Biochem.
 
33
,
93
102
.
[62]
Castro
M.S.
Steudler
P.A.
Melillo
J.M.
(
1995
)
Factors controlling atmospheric methane consumption by temperate forest soils
.
Global Biogeochem. Cycles
 
9
,
1
10
.
[63]
Menyailo
O.V.
Hungate
B.A.
(
2003
)
Interactive effects of tree species and soil moisture on methane consumption
.
Soil Biol. Biochem.
 
35
,
625
628
.
[64]
Borken
W.
Beese
F.
Lamersdorf
N.
(
2002
)
Long-term reduction in nitrogen and proton inputs did not affect atmospheric uptake and nitrous oxide emission from a German spruce forest soil
.
Soil Biol. Biochem.
 
34
,
1815
1819
.
[65]
Dunfield
P.F.
Topp
E.
Archambault
C.
Knowles
R.
(
1995
)
Effetc of nitrogen fertilizers and moisture content on CH4 and N2O fluxes in a humisol: measurements in the field and intact soil cores
.
Biogeochem.
 
29
,
199
222
.
[66]
Glatzel
S.
Stahr
K.
(
2001
)
Methane and nitrous oxide exchange in differently fertilised grassland in southern Germany
.
Plant Soil
 
231
,
21
35
.
[67]
Lessard
R.
Rochette
P.
Gregorich
E.G.
Desjardins
R.L.
Pattey
E.
(
1997
)
CH4 fluxes from a soil amended with dairy cattle manure and ammonium nitrate
.
Can. J. Soil Sci.
 
77
,
179
186
.
[68]
Ruser
R.
Flessa
H.
Schilling
R.
Steindle
H.
Beese
F.
(
1998
)
Soil compaction and fertilisation effects on nitrous oxide and methane fluxes in potato fields
.
Soil Sci. Soc. Am. J.
 
62
,
1587
1595
.
[69]
Kamann
C.
Grünhage
L.
Jäger
H.J.
Wachinger
G.
(
2001
)
Methane fluxes from differentially managed grassland study plots: the important role of CH4 oxidation in grassland with a high potential for CH4 production
.
Environ. Pollut.
 
115
,
261
273
.
[70]
Dobbie
K.E.
Smith
K.A.
(
1996
)
Comparison of CH4 oxidation rates in woodland, arable and set aside soils
.
Soil Biol. Biochem.
 
28
,
1357
1365
.
[71]
Nykänen
H.
Vasander
H.
Huttunen
J.T.
Martikainen
P.J.
(
2002
)
Effect of experimental nitrogen load on methane and nitrous oxide fluxes on ombrotrophic boreal peatland
.
Plant Soil
 
242
,
147
155
.
[72]
Moosavi
S.C.
Crill
P.M.
(
1998
)
CH4 oxidation by tundra wetlands as measured by a selective inhibitor technique
.
J. Geophys. Res.
 
27
,
29093
29106
.
[73]
Boon
P.I.
Lee
K.
(
1997
)
Methane oxidation in sediments of a floodplain wetland in south-eastern Australia
.
Lett. Appl. Microbiol.
 
25
,
138
142
.
[74]
Schnell
S.
King
G.M.
(
1995
)
Mechanistic analysis of ammonium inhibition of atmospheric methane consumption in forest soil
.
Appl. Environ. Microbiol.
 
60
,
3514
3521
.
[75]
Eller
G.
Frenzel
P.
(
2000
)
Changes in activity and community structure of methane-oxidizing bacteria over the growth period of rice
.
Appl. Environ. Microbiol.
 
67
,
2395
2403
.
[76]
Krüger
M.
Eller
G.
Conrad
R.
Frenzel
P.
(
2002
)
Seasonal variation in pathways of CH4 oxidation in rice fields determined by stable carbon isotopes and specific inhibitors
.
Global Change Biol.
 
8
,
265
280
.
[77]
Dan
J.
Krüger
M.
Frenzel
P.
Conrad
R.
(
2001
)
Effect of a late season urea fertilization on methane emission from a rice field in Italy
.
Agric. Ecosyst. Environ.
 
69
,
69
80
.
[78]
Sitaula
B.K.
Bakken
L.R.
(
1993
)
Nitrous oxide release from spruce forest soil: relationships with nitrification, methane uptake, temperature, moisture and fertilisation
.
Soil Biol. Biochem.
 
25
,
1415
1421
.
[79]
Kruse
C.W.
Iversen
N.
(
1995
)
Effect of plant succession, ploughing, and fertilisation on the microbiological oxidation of atmospheric methane in a heath land soil
.
FEMS Microbiol. Ecol.
 
18
,
121
128
.
[80]
Goldman
M.B.
Groffman
P.M.
Pouyat
R.V.
McDonnel
M.J.
Pickett
T.A.
(
1995
)
CH4 uptake and N availability in forest soils along an urban to rural gradient
.
Soil Biol. Biochem.
 
27
,
281
286
.
[81]
Bender
M.
Conrad
R.
(
1995
)
Effect of CH4 concentrations and soil conditions on the induction of CH4 oxidation activity
.
Soil Biol. Biochem.
 
27
,
1517
1527
.
[82]
Boeckx
P.
VanCleemput
O.
Villaralvo
I.
(
1997
)
Methane oxidation in soils with different textures and land use
.
Nutr. Cycl. Agroecosyst.
 
49
,
91
95
.
[83]
Willison
T.W.
O'Flaherty
M.S.
Tlustos
P.
Goulding
K.W.T.
Powlson
D.S.
(
1997
)
Variations in microbial populations in soils with different methane uptake rates
.
Nutr. Cycl. Agroecosyst.
 
49
,
85
90
.
[84]
Priemé
A.
Christensen
S.
Dobbie
K.E.
Smith
K.A.
(
1997
)
Slow increase in rate of methane oxidation in soils with time following land use change from arable agriculture to woodland
.
Soil Biol. Biochem.
 
29
,
1269
1273
.
[85]
Rigler
E.
Zechmeister-Boltenstern
S.
(
1999
)
Oxidation of ethylene and methane in forest soils-effect of CO2 and mineral nitrogen
.
Geoderma
 
90
,
147
159
.
[86]
de Visscher
A.
Thomas
D.
Boeckx
P.
Van Cleemput
O.
(
1999
)
Methane oxidation in stimulated landfill cover soil environments
.
Environ. Sci. Technol.
 
33
,
1854
1859
.
[87]
Börjesson
B.
Nohrstedt
H.Ö.
(
2000
)
Fast recovery of atmospheric methane consumption in a Swedish forest soil after single-shot N-fertilisation
.
Forest Ecol. Manage.
 
134
,
83
88
.
[88]
Chan
A.S.K.
Parkin
T.B.
(
2001
)
Methane oxidation and production activity in soils from natural and agricultural ecosystems
.
J. Environ. Qual.
 
30
,
1896
1903
.
[89]
Hellebrand
H.J.
Kern
J.
Scholz
V.
(
2003
)
Long-term studies on greenhouse gas fluxes during cultivation of energy crops on sandy soils
.
Atmos. Environ.
 
37
,
1635
1644
.
[90]
Anthony
C.
(
1982
)
The Biochemistry of Methylotrophs
 .
Academic Press
,
London
.
[91]
Megraw
S.R.
Knowles
R.
(
1987
)
Active methanotrophs suppress nitrification in a humisol
.
Biol. Fertil. Soils
 
4
,
205
212
.
[92]
Murrell
J.C.
Dalton
H.
(
1983
)
Nitrogen fixation in obligate methanotrophs
.
J. Gen. Microbiol.
 
129
,
3481
3486
.
[93]
Aumann
A.J.
Speake
C.C.
Lidstrom
M.E.
(
2001
)
nifH Sequences and nitrogen fixation in type I and type II methanotrophs
.
Appl. Environ. Microbiol.
 
67
,
4009
4016
.
[94]
Park
S.
Shah
N.H.
Taylor
R.T.
Droege
M.W.
(
1992
)
Batch cultivation of Methylosinus trichosporium OB3b: II. Production of particulate methane monooxygenase
.
Biotechnol. Bioeng.
 
40
,
151
157
.
[95]
Sullivan
J.P.
Dickinson
D.
Chase
H.A.
(
1998
)
Methanotrophs, Methylosinus trichosporium OB3b, and their application to bioremediation
.
CRC Microbiol.
 
24
,
335
373
.
[96]
Merrick
M.J.
Edwards
R.A.
(
1995
)
Nitrogen control in bacteria
.
Microbiol. Rev.
 
59
,
604
622
.
[97]
Murrell
J.C.
(
1982
)
The rapid switch-off of nitrogenase activity in obligate methane-oxidizing bacteria
.
Arch. Microbiol.
 
150
,
489
495
.
[98]
Anthony
C.
(
1978
)
The prediction of growth yields in methylotrophs
.
J. Gen. Microbiol.
 
104
,
91
104
.
[99]
Jones
R.D.
Morita
R.Y.
(
1983
)
Methane oxidation by Nitrosococcus oceanus and Nitrosomonas europaea
.
Appl. Environ. Microbiol.
 
45
,
401
410
.
[100]
Jiang
Q.Q.
Bakken
L.R.
(
1999
)
Nitrous oxide production and methane oxidation by different ammonia-oxidising bacteria
.
Appl. Environ. Microbiol.
 
65
,
2679
2684
.
[101]
Popp
T.J.
Chanton
J.P.
Whiting
G.J.
Grant
N.
(
2000
)
Evaluation of methane oxidation in the rhizosphere of a Carex dominated fen in north central Alberta, Canada
.
Biogeochemistry
 
51
,
259
281
.
[102]
Granberg
G.
Sundh
I.
Svensson
B.H.
Nilsson
M.
(
2001
)
Effects of temperature, and nitrogen and sulfur deposition, on methane emission from a boreal mire
.
Ecology
 
82
,
1982
1998
.
[103]
Jensen
S.
Priemé
A.
Bakken
L.
(
1998
)
Methanol improves methane uptake in starved methanotrophic microorganisms
.
Appl. Environ. Microbiol.
 
64
,
1143
1146
.
[104]
West
A.E.
Schmidt
S.K.
(
1999
)
Acetate stimulates atmospheric CH4 oxidation by an alpine tundra soil
.
Soil Biol. Biochem.
 
31
,
1649
1655
.
[105]
Groot
T.T.
Van Bodegom
P.M.
Harren
F.J.M.
Meijer
H.A.J.
(
2003
)
Quantification of methane oxidation in the rice rhizosphere using 13C-labelled methane
.
Biogeochemistry
 
64
,
355
372
.
[106]
van der Nat
F.J.W.A.
Middelburg
J.J.
(
1998
)
Seasonal variation in methane oxidation by the rhizosphere of Phragmites australis and Scirpus lacustris
.
Aquat. Bot.
 
61
,
95
110
.
[107]
Updegraff
K.
Bridgham
S.D.
Pastor
J.
Weishampel
P.
Harth
C.
(
2001
)
Response of CO2 and CH4 emissions from peat lands to warming and water table manipulation
.
Ecol. Appl.
 
11
,
311
326
.
[108]
Nold
S.C.
Boschker
H.T.S.
Pel
R.
Laanbroek
H.J.
(
1999
)
Ammonium addition inhibits 13C-methane incorporation into methanotroph lipids in a freshwater sediment
.
FEMS Microbiol. Ecol.
 
29
,
81
89
.
[109]
Radajewski
S.
Ineson
P.
Parekh
N.R.
Murrell
J.C.
(
2000
)
Stable-isotope probing as a tool in microbial ecology
.
Nature
 
403
,
646
649
.
[110]
Manefield
M.
Whiteley
A.S.
Griffiths
R.I.
Bailey
M.
(
2002
)
RNA stable isotope probing, a novel means of linking microbial community function to phylogeny
.
Appl. Environ. Microbiol.
 
68
,
5367
5373
.
[111]
Benstead
J.
King
G.M.
Williams
H.G.
(
1998
)
Methanol promotes atmospheric methane oxidation by methanotrophic cultures and soils
.
Appl. Environ. Microbiol.
 
64
,
1091
1098
.