1. Carbon storage or sequestration?
Storage (as defined by Millard, 1988) involves C being held in a pool, with the potential to be reused subsequently for growth or maintenance of another. In contrast, sequestration represents a metabolic dead end, with the C no longer physiologically active and so not affecting metabolism. So do trees store C? The short answer is yes, but much of the nonstructural soluble carbohydrates (NSC), such as starch, that are accumulated in woody tissues are sequestered and so probably not reusable by the tree. Seasonal fluctuations in the concentrations of NSC such as starch have also been found in many species, especially in seedlings (Gansert & Sprick, 1998). The ability to store C (and nutrients such as N and P) is important in allowing trees to recover from disturbances such as defoliation, and the relative ability to store C has been associated with seedling survival and growth in temperate (Kobe, 1997), tropical (Gleason & Ares, 2004) and boreal species (Kagawa et al., 2006a). C remobilization in deciduous trees has also been demonstrated as an important process during bud burst (Maurel et al., 2004), providing up to about 40% of the C used for new leaf growth in a boreal species with a short growing season (Kagawa et al., 2006a) and contributing to early wood formation (Kagawa et al., 2006b). C remobilization can also be important for recovery from winter embolism (Ameglio et al., 2004).
Trees certainly accumulate large pools of NSC. Hoch et al. (2003a) estimated that in a temperate forest the concentrations of NSC found in mature trees would allow the whole leaf canopy to be replaced four times. Even when tree growth is constrained to a short season by temperature (e.g. at the tree line) the build-up of NSC suggests that C availability is not a limitation to growth (Hoch & Körner, 2003). Furthermore, NSC pools in trees are never fully depleted in the same way that N storage pools are (e.g. Hoch et al., 2003a). Even girdling the phloem does not deplete NSC in roots completely (Jordan & Habib, 1996; Bhupinderpal-Singh et al., 2003), suggesting that roots contain pools of NSC which are not utilized, even when the supply of current assimilates is cut. So is the build-up of NSC in leaves, trunks and roots of trees C storage or sequestration? It seems to be partly a mechanism to avoid down-regulation of photosynthesis, sequestering C that cannot be reused by the tree. This is a mechanism comparable to arginine accumulation in conifers in response to excessive N deposition (Näsholm, 1994).
In comparison, it is well established that trees store nutrients (Millard, 1996) and that, particularly for N, remobilization of stored resources can provide the first source used for growth during flushing and the majority used for growth above ground each year (Dyckmans & Flessa, 2001; Millard et al., 2001; Carswell et al., 2003). N remobilization from storage is source-driven (Millard et al., 2001). In consequence, in contrast to C in NSC pools, all the stored N is depleted during remobilization, as has been seen in temperate species by, for example, the complete disappearance of bark storage proteins during the spring (Wetzel et al., 1989; Cooke & Weih, 2005).
2. Leaf senescence
Nutrient withdrawal from leaves during senescence is often very efficient and in deciduous trees plays an important role in their internal cycling of N and P (Millard, 1996). The remaining nutrient fraction returns to the soil as litter, with the mass and nutrient content of litter produced being the primary determinant of nutrient availability in forest soils (Prescott, 2002). The fraction of mineral nutrients withdrawn from leaves before their abscission is variable with differences associated with leaf life span, overall leaf nutrient content, soil fertility (Niinemets & Tamm, 2005) and environmental conditions such as drought (Del Arco et al., 1991). In a review comparing over 100 deciduous shrubs and trees, Aerts (1996) reported that 54 ± 16% of the N and 50 ± 20% of the P was withdrawn from a leaf, with resorption efficiencies usually lower for trees growing on sites with higher fertility. Van Heerwaarden et al. (2003) suggested that these values could be increased by a further 6–10% when accounting for changes in leaf mass and areas during senescence, the implication being that the majority of leaf nutrients were reused by the tree. In addition, trees have also evolved pigment systems (such as anthocyanins) to protect leaves from radiation damage and so facilitate nutrient recovery during senescence (Hoch et al., 2003b). One of the first events to occur during senescence is a turnover of Rubisco and export of N (as amino acids) from the leaf. This selective breakdown of Rubisco can lead to an approximately 80-fold greater rate of turnover than of other enzymes of the Calvin cycle (Crafts-Brander et al., 1990), thus emphasizing the role of Rubisco as a N storage protein.
In contrast to nutrient withdrawal from senescing leaves, much of the soluble C present at the end of the summer remains in the abscised leaf. Indeed, leaf senescence is now thought to be triggered by the accumulation of both soluble sugars and starch in the leaf (Ono et al., 2001; Pourtau et al., 2004) and it is the water-soluble C content (i.e. sugars) that may determine initial litter decomposition rate (Allison & Vitousek, 2004). In the longer term at the community scale, leaf content in secondary C compounds (such as lignin, polyphenols and tannins) is negatively correlated with litter decomposition rate. As slow-growing trees tend to produce leaves rich in such compounds, the feedback between leaf quality and decomposition rate contributes to maintaining lower levels of soil fertility and is thought to explain the dominance of such trees in their habitat (Aerts, 1999; Cornelissen et al., 1999). Hence C in leaf litter, in the form of complex secondary compounds or sugars, plays an important part in determining the rate of soil nutrient cycling. C investment in secondary C compounds is usually regarded as a means to protect the key organs in C acquisition against herbivory. However, many evergreen tree species store N in their leaves and, in their case, antiherbivory compounds may serve primarily to protect the nutrients stored within the leaves (Millard et al., 2001). Thus trees tend to protect and then reuse the nutrients held in their leaves, while allowing much of the soluble leaf C to abscise with the leaf. Taken together, these physiological and structural plant mechanisms suggest an evolutionary pressure led by nutrient rather than C limitation.
3. The importance of C allocation to roots and associated microbes
In addition to the C and nutrients returned to the soil as litter, trees allocate much of their C below ground. Giardina et al. (2005) strikingly summarized the importance of below-ground C allocation: half of the 120 Pg C fixed annually by terrestrial plants is allocated below ground, with tree-based ecosystems accounting for most of this flux, amounting to 20 times the annual release of C by combustion of fossil fuels. Below-ground C allocation has been estimated as being between 35 and 50% of net primary production (NPP) in a tropical forest (Giardina et al., 2003), and as much as 73% in Douglas fir (Fogel & Hunt, 1983) and black spruce (Ruess et al., 2003) forests. Figure 2 shows the pathways for C input to the soil: from litter (pathway 1); by transfer from roots to mycorrhizal fungi (pathway 2); directly into the soil as mycorrhizodeposits (pathway 3) or secreted enzymes (pathway 4); and through grazing by soil fauna (pathway 5). The profound impact of this C (particularly through pathway 2 in Fig. 2) on the functioning of soil microbial communities has been demonstrated in large-scale field experiments. Girdling trees has been used to stop the flow of photosynthates to roots (Högberg et al., 2001), thereby reducing soil respiration. In a tropical eucalyptus forest, girdling reduced soil respiration by 16–24% (Binkley et al., 2006), despite tree canopies remaining intact for 3 months and live fine root biomass showing no decrease for at least a further 2 months, indicating C storage capable of sustaining root maintenance and respiration. Girdling lowered soil respiration by more than 50% in forests in northern Sweden (Bhupinderpal-Singh et al., 2003) and Germany (Subke et al., 2004) and 31–44% in Colorado (Scott-Denton et al., 2006). Furthermore, there was an average decline of 32% of the soil microbial biomass within 3 months at the Swedish site, attributed by Högberg & Högberg (2002) to a loss of the extra radical ectomycorrhizal mycelium.
The formation of associations with mycorrhizal fungi is ubiquitous in the plant kingdom. Roots of forest trees are heavily colonized by ectotrophic (ectomycorrhizal, EM) and/or endotrophic (arbuscular mycorrhizal, AM) fungi. For example, Ruess et al. (2003) found 100% of first-order roots to be EM in three mature black spruce forests, whilst Adams et al. (2006) reported colonization rates by EM fungi of between 76 and 100% of root length in eucalyptus forests. Gehring & Connell (2006) reported a colonization rate of up to 61% of root length of seedlings in tropical and subtropical rain forests. A meta-analysis of data for root colonization by AM fungi found an average (across all plant species) of 36 ± 10, 23 ± 3 and 24 ± 8% root length in tropical, temperate and boreal forests, respectively (Treseder & Cross, 2006). Boreal and temperate forest trees predominantly associate with EM species, whilst the majority of forest trees growing in tropical climates have AM symbioses. There are, of course, exceptions to this crude classification, with a minority of species forming associations with EM fungi in tropical forests (Alexander, 2006) or AM fungi in temperate forests (e.g. sugar maple, sweetgum, some eucalypts and poplars). When co-occurring in the same ecosystem, EM and AM fungi have been shown to occupy different niches, with a greater abundance of EM species in organic soil horizons, and with AM species predominating in mineral horizons (Neville et al., 2002; but see Moyersoen et al., 1998). Some species also exhibit successional mycorrhizal associations, often with an increase in the abundance of EM and a decrease of AM fungi as the trees age, as reported by Adams et al. (2006) in their comparison of field-grown seedlings with mature Eucalyptus grandis trees. Similarly, Egerton-Warburton & Allen (2001) found a shift in the relative abundance of AM and EM fungi in oak trees along a 1–30 yr age sequence (presumably as tree carbohydrate storage capacity increased) and suggested that C supply to the roots partly controls succession from AM to EM, adding weight to the view that EM and AM fungi exhibit different C sink strengths (Lynch & Whipps, 1990). So how much C is actually transferred from roots to EM and AM mycorrhizal fungi?
Microcosm studies of EM conifers have shown that up to 30% of total photoassimilates are transferred to the fungal partner (reviewed by Soderström, 2002). Soil C balance calculations for a field site in a northern hardwood forest estimated that 17% of total below-ground C allocation was to mycorrhizal fungi (in a mixed AM and EM community) and exudation (Fahey et al., 2005). However, to our knowledge, there has been no direct quantification in situ of the total C flux between tree roots and their mycorrhizal partners. The main impediment to direct measurements of C allocation to mycorrhizal fungi is a lack of adequate techniques. C transfer to mycorrhizal fungi can be considered as two fluxes: C exchange at the root–fungus interface (flux 2α, Fig. 2) and C allocation between intra- and extraradical hyphae (flux 2β, Fig. 2). The quantification in situ of C transfer at the plant–fungus interface (flux 2α, Fig. 2) is virtually impossible, without either full characterization of transport systems and specific compounds involved in resource transfer, or identification and isolation of a pool of fungal/plant specific compounds with relatively rapid turnover (such as nucleic acids) which could be used as markers of C incorporation. This second approach, together with the rapid development of molecular techniques, may allow study of both function and structure of the mycorrhizal community intercepting plant C (Johnson et al., 2005a). This would be particularly relevant, given emerging evidence that the structure of the EM community is affected by treatments changing C supply below ground (Fransson et al., 2001; Parrent et al., 2006). Root-excluding mesh cores, allowing ingrowth of mycorrhizal mycelia, combined with isotopic C labelling, have been used to quantify the flux of C to extraradical mycelia of AM fungi in grassland ecosystems (Johnson et al., 2002), hence describing the flux of C transfer from intraradical to the network of extra radical hyphae (flux 2β, Fig. 2) which can extend centimetres (AM) to metres (EM) into the soil and link plant roots together (Selosse et al., 2006). Such an approach has been used in forest ecosystems to measure fungal biomass, but not for a full C budget, because C loss through fungal respiration was not measured (Godbold et al., 2006). Another major limitation to establishing an accurate C budget of the mycorrhizal community in forest ecosystems is our ability to measure mycorrhizal biomass accurately; particularly to distinguish extraradical mycorrhizal hyphae from mycelia of saprotrophic species (Wallander, 2006). Extraradical hyphae may contribute substantially to soil C stock. It has been estimated that over 60% of the accumulation of soil organic C at a poplar FACE site was derived from external mycorrhizal hyphae, twice that from leaf litter (Godbold et al., 2006). In terms of litter production (flux 1δ, Fig. 2), such large amounts of fungal biomass in soils may have important implications for rates of soil organic matter (SOM) decomposition, particularly when comparing EM- and AM-dominated forests. The two types involve fungal species with fundamentally different morphology and chemical composition, which potentially affects their palatability to soil fauna. This will have inevitable consequences on the substantial amounts of C (Johnson et al., 2005b) going from mycorrhizal fungi to the soil food web (flux 5δ, Fig. 2). Hyphal turnover rates also differ between AM and EM (Staddon et al., 2003; Treseder et al., 2004), and AM produce substances (e.g. glomalin) which can stabilize soil aggregates (Rillig & Mummey, 2006). The chemical and morphological characteristics of mycorrhizal hyphae may also affect the fine root decomposition rate. When investigating the decomposition rate of EM vs nonmycorrhizal pine roots, Langley et al. (2006) showed that EM roots lost only one-third of the C of nonmycorrhizal roots. Langley & Hungate (2003) hypothesized that mycorrhizal type may ‘substantially influence fine root decomposition and soil carbon processing rate’.
So why do trees allocate such a substantial proportion of their C below ground and specifically to mycorrhizal fungi? The common view is that through their mycorrhizal partners, trees access nutrients otherwise unavailable for direct root uptake (such as complex forms of organic P and N), or present in insufficient quantities in the vicinity of the root (e.g. orthophosphate; Hinsinger, 2001). For example, using a microcosm system, Brandes et al. (1998) calculated uptakes of 73% of tree N and 76% of P via the EM fungus Paxillus involutus colonizing the roots of Norway spruce. In addition, van Breemen et al., 2000) suggested that external hyphae play a substantial role in nutrient return to the tree, via mineral weathering (Fig. 2, pathway C). Read (1991) hypothesized that there is a link between plant biome, mycorrhizal type and overall SOM amount, soil pH and N : P status. Read's hypothesis, later reinforced by Read & Perez-Moreno (2003), explains both the broad relative distribution of both mycorrhizal types and the body of evidence showing that AM fungi contribute predominantly to their host P requirements (but see Hodge et al., 2001) through enhancement of phosphorus uptake, while EM fungi have a greater capability to mobilize N and P from SOM. In boreal forests, this could satisfy a considerable proportion of the annual nutrient requirement of the trees (Read & Perez-Moreno, 2003). Recently, understanding the functional difference between AM and EM might have been taken a step further. Lindahl et al. (2007) demonstrated that saprotrophic fungi were mainly found in the soil horizons where primary decomposition of recent leaf litter occurs, whilst EM fungal taxa dominated the underlying horizon where N was mobilized from partly decomposed litter. Their study added weight to the view that AM forests (tropical) have an extravagant nutrient cycle relying heavily on N mineralization, whereas EM forests (temperate/boreal) have a conservative nutrient cycle where trees access organic N via EM fungi, bypassing mineralization processes by free-living microbes (Chapman et al., 2006).
It may be assumed that the ‘C cost’ of nutrient acquisition through symbiotic partners is higher than that of acquiring nutrients through direct root uptake from the soil solution. Does this imply that the C demand by the symbiotic partner leads to C limitation of tree growth? The consensus view is that this is not the case. Mycorrhizal infection has been shown to cause an up-regulation of photosynthesis in young EM trees in microcosms (Loewe et al., 2000) and pots (Wright et al., 2000) and young AM trees (Lovelock et al., 1997) in pots, suggesting that when the mycorrhizal associations impose an extra C demand on the tree, it can be met by increasing C assimilation. The initial presence of the fungi may lead to a growth depression for trees under conditions of high soil fertility, as shown in field experiments in highly managed agrosystems (Graham & Eissenstat, 1998); presumably as the benefit to the plant of the increased nutrient availability conferred by the fungus diminishes. However, such levels of soil fertility are not found in forest ecosystems. To assess the C cost to the tree of nutrient acquisition via mycorrhizal fungi in situ, we need to understand the regulation of C exchange at the plant–root interface by both biological (e.g. plant C and nutrient status) and environmental factors (e.g. nutrient availability in the mycorrhizosphere, temperature). This is assuming that the amount of C passed on to its symbionts (2α, Fig. 2) is mainly controlled by the plant itself. However such a phytocentric view is now being questioned, and it is argued that control of C use and allocation (flux 2β, Fig. 2) by mycorrhizal partner(s) also need to be considered (Lindahl et al., 2002; Staddon, 2005), especially given the extent of extraradical mycorrhizal hyphae produced by EM fungi. In a meta-analysis comparing the response ratio of AM/EM plants (not exclusively trees) and mycorrhizal fungi to elevated CO2, Alberton et al. (2005) reported evidence ‘for the mycocentric view in EM, but not in AM systems’.
In addition to transferring C to their root symbionts, trees lose C from their roots as rhizodeposits (pathway 3, Fig. 2). Rhizodeposition involves a wide range of compounds (Grayston et al., 1997), including exudation of low-molecular-weight substances such as sugars, organic acids and amino acids. The availability of these C substrates is considered the factor most limiting to the growth of free-living soil microbes (Wardle, 1992), explaining the greater microbial activity in rhizosphere compared with bulk soil. There are few reliable quantitative estimates of the flux of C entering the soil as exudates. A review of whole-plant 14C-labelling studies performed in soil on a wide range of plant species suggested that exudation accounted for 5–10% of net C assimilation (Farrar et al., 2003), although Jones et al. (2004) highlighted a possible overestimation resulting from methodological bias and suggested that a true estimate of root exudation was likely to be only 2–4% of net fixed C. Under a variety of stress conditions (including nutrient or water stress) the C flux from roots as exudates is increased, mainly because of damage to membranes or disruption of normal cell metabolism (Neumann & Römheld, 2001). As noted by Jones et al. (2004), the vast majority of studies quantifying root exudates have not considered the quantitative or qualitative impact of the (ubiquitous) mycorrhizal fungi colonizing roots (Fig. 2, pathway 3β), despite numerous studies demonstrating exudation of hydrolytic enzymes by mycorrhizal mycelium (mainly EM fungi). Recently Phillips & Fahey (2006) calculated that microbial biomass, C/N mineralization rates and phosphatase activity were 25–30% higher in the rhizosphere than in bulk soil under EM tree species. Under AM trees, rhizosphere and bulk soil differed by only 10–12% for similar parameters. Tree C inputs to soil as rhizodeposits can also act as primers for the degradation of existing SOM (Fontaine et al., 2004; Hoosbeek et al., 2004) through both abiotic and biotic mechanisms (reviewed by Kuzyakov et al., 2000). However, the trade-off between C priming and nutrient availability, on the one hand, and tree growth on the other remains largely unknown. Under elevated CO2, exudation is thought to be increased on a per-plant basis via stimulation of root growth, although experimental evidence is still lacking.
Taken together, C transferred from trees to root symbionts and rhizodeposition account for a significant proportion of net C assimilation. The tree benefits from this C loss in terms of microbial activity and the consequent nutrient acquisition and cycling through the turnover of SOM. These complex interactions between the tree and the soil suggest that C availability is not the primary resource limitation for tree growth. If C were limiting tree growth, it would only be through C limitation to nutrient cycling in forest soils.