Do tree species influence soil carbon stocks in temperate and boreal forests?

Do tree species influence soil carbon stocks in temperate and boreal forests?

Forest Ecology and Management xxx (2013) xxx–xxx Contents lists available at SciVerse ScienceDirect Forest Ecology and Management journal homepage: ...

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Forest Ecology and Management xxx (2013) xxx–xxx

Contents lists available at SciVerse ScienceDirect

Forest Ecology and Management journal homepage:

Do tree species influence soil carbon stocks in temperate and boreal forests? Lars Vesterdal a,⇑, Nicholas Clarke b, Bjarni D. Sigurdsson c, Per Gundersen a a

Department of Geosciences and Natural Resource Management, University of Copenhagen, Rolighedsvej 23, DK-1958 Frederiksberg C, Denmark Norwegian Forest and Landscape Institute, P.O. Box 115, N-1431 Ås, Norway c Agricultural University of Iceland, Hvanneyri, IS-311 Borgarnes, Iceland b

a r t i c l e

i n f o

Article history: Available online xxxx Keywords: Tree species Carbon sequestration Soil organic carbon Forest floor Mineral soil Review

a b s t r a c t Information on tree species effects on soil organic carbon (SOC) stocks is scattered and there have been few attempts to synthesize results for forest floor and mineral soil C pools. We reviewed and synthesized current knowledge of tree species effects on SOC stocks in temperate and boreal forests based on common garden, retrospective paired stand and retrospective single-tree studies. There was evidence of consistent tree species effects on SOC stocks. Effects were clearest for forest floor C stocks (23 of 24 studies) with consistent differences for tree genera common to European and North American temperate and boreal forests. Support for generalization of tree species effects on mineral soil C stocks was more limited, but significant effects were found in 13 of 22 studies that measured mineral soil C. Proportional differences in forest floor and mineral soil C stocks among tree species suggested that C stocks can be increased by 200–500% in forest floors and by 40–50% in top mineral soil by tree species change. However, these proportional differences within forest floors and mineral soils are not always additive: the C distribution between forest floor and mineral soil rather than total C stock tends to differ among tree species within temperate forests. This suggests that some species may be better engineers for sequestration of C in stable form in the mineral soil, but it is unclear whether the key mechanism is root litter input or macrofauna activity. Tree species effects on SOC in targeted experiments were most consistent with results from large-scale inventories for forest floor C stocks whereas mineral soil C stocks appeared to be stronger influenced by soil type or climate than by tree species at regional or national scales. Although little studied, there are indications that higher tree species diversity could lead to higher SOC stocks but the role of tree species diversity per se vs. species identity effects needs to be disentangled in rigorous experimental designs. For targeted use of tree species to sequester soil C we must identify the processes related to C input and output, particularly belowground, that control SOC stock differences. We should also study forms and stability of C along with bulk C stocks to assess whether certain broadleaves store C in more stable form. Joint cooperation is needed to support syntheses and process-oriented work on tree species and SOC, e.g. through an international network of common garden experiments. Ó 2013 Elsevier B.V. All rights reserved.

1. Why do we need to know about tree species effects? Vegetation type, including dominant tree species, has since long been recognized among the important soil forming factors (Jenny, 1994; Hobbie, 1992). In the temperate region and also in parts of the boreal region, the present tree species composition in forests is often a result of past forest management decisions. The selection of tree species has mainly been driven by the objective to optimize productivity of stemwood and demands for wood of certain tree species. However, with the increasing emphasis on ecosystem services other than wood from forestry and increasing focus on ⇑ Corresponding author. Tel.: +45 35331672; fax: +45 35331517. E-mail addresses: [email protected] (L. Vesterdal), [email protected] (N. Clarke), [email protected] (B.D. Sigurdsson), [email protected] (P. Gundersen).

adaptation to climate change (Lindner et al., 2010) there is a need to provide a wider basis for informed decisions regarding tree species selection. Such informed decisions include the potential to sequester soil organic carbon (SOC) through selection of a wider range of tree species. Climate change and associated disturbances are also expected to influence the species composition in temperate and boreal forests (Allen et al., 2010; Lindner et al., 2010) and could be important in regions where natural disturbances control regeneration and C cyling processes. Yet we know little about the consequences of climate change induced dynamics in tree species distribution on the capacity of soils to store carbon (Jones et al., 2009). Recently, the role of forest SOC stocks and their dynamics for mitigation of greenhouse gases (GHG) has highlighted the need for more knowledge on tree species effects (Jandl et al., 2007). For-

0378-1127/$ - see front matter Ó 2013 Elsevier B.V. All rights reserved.

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est management, including a change in tree species, and afforestation of cropland and grassland are accepted measures for mitigation of atmospheric CO2 in national GHG budgets. However, quantitative estimates of tree species effects on SOC stocks are still scarce and the scientific basis for targeted use of tree species to maximize sequestration of SOC following afforestation is also limited to relatively few studies (Vesterdal et al., 2002; Perez-Cruzado et al., 2012). Several other biotic and abiotic factors, such as soil type, soil water status and climate, influence soil C pools (Callesen et al., 2003; Baritz et al., 2010). Since different tree species are not randomly distributed in natural forest ecosystems, but indeed tend to follow certain gradients in climate, soil types, soil water status, forest successional stage or other abiotic factors that also directly affect the soil C-status, it is doubtful if a direct comparison between C stocks of natural forests will say anything about the true ‘‘species effects’’ on the present soil C. Experimental plots that limit the influence of such confounding factors in natural or old-growth forests are, however, rare (Binkley, 1995). The influence of tree species on forest soil properties has for a long time been studied by ecologists but the focus of these studies has not until recently been directed toward effects on SOC stocks and their dynamics for mitigation of greenhouse gases. The Danish silviculturalist P.E. Müller was among the first to note the differences in humus accumulation and turnover under different forest vegetation, and his work introduced the Danish terms mull and mor to the international pedological vocabulary (Müller, 1887). Later, the interest in tree species effects on soils was mainly focused on soil fertility parameters and pedological influence of different tree species and possible environmental problems following e.g. deposition of nitrogen and heavy metals (Zinke, 1962; Binkley, 1995; Binkley and Giardina, 1998). Previous studies suggested that tree species planted within the same sites may differ in SOC (Finzi et al., 1998a; Mareschal et al., 2010), but the main effect has been suggested to occur in less protected forest floor C pools (Vesterdal et al., 2008). However, it is highly uncertain which processes are responsible for differences in soil C stocks. Differences in input (Díaz-Pinés et al., 2011) as well as decomposition rates (Vesterdal et al., 2008; Hansen et al., 2009) have been suggested as the most likely explanation for tree species influence on soil C stocks. For targeted use of tree species to sequester SOC, we therefore urgently need to identify which processes control C stock differences and to study forms and stability of C within stocks of bulk C. There have been very limited attempts to synthesize and review tree species effects on SOC stocks in light of the possibility to sequester C in soils by management decisions to change the dominant tree species. In a literature review of forest management effects on SOC storage by Johnson (1992), ‘‘species change’’ effects were briefly addressed as one of several possible management parameters. He concluded that the effect of tree species on soil C was often significant, but inconsistent. This result was probably partly caused by the limited number of tree species studies received on his letter request. Binkley and Giardina (1998) also mentioned that species effects on SOC stocks lack experimental testing within the same site. They synthesized global forest floor mass differences among tree species and found that masses commonly differed by 20% within common garden plots (i.e. a factor 1.2), but differences were often larger, i.e. up to 5-fold. Since the publication of these two reviews, a number of experimental studies on tree species and SOC have been published. This inspired us to synthesize the current available evidence of tree species effects on SOC. The aim of this study was thus to synthesize current knowledge of tree species effects on SOC stocks in temperate and boreal forests. We limited our analysis to these specific regions to enable

detection of possible patterns among common tree species and genera. We specifically addressed C stocks in forest floor and mineral soil and their relative contribution to total SOC stocks as well as evidence of the most influential input and output processes in control of tree species effects on SOC. In order to ensure minimal influence of site-related confounding factors we only focused on evidence from multi-species comparisons within experimental designs with ‘‘similar’’ soil conditions. This information will enable prediction of the direction of soil C stock change on a decadal time scale following a change in tree species. We also synthesized studies on tree species effects on soil C stocks from larger-scale inventories to see if similar patterns in tree species effects occurred across gradients in site properties. Lastly we identified areas for future research to disentangle mechanisms behind tree species effects on soil C stocks.

2. Experimental designs used to study tree species effects on SOC Many different designs have been used over time to study tree species effects, e.g. structured observations of soil organic matter under different vegetation types by early scientists like Müller (1887), retrospective studies under single trees (e.g. Zinke, 1962; Dijkstra and Fitzhugh, 2003), retrospective studies of paired adjacent stands of different species (e.g. Mergen and Malcolm, 1955; Laganière et al., 2012) and planned multispecies common garden experiments with different tree species planted in randomized plots at the same time on similar soil (e.g. Challinor, 1968; Reich et al., 2005). We use the term ‘retrospective’ in connection with single tree and paired stand studies since the experimenter often did not set these designs, but expected important conditions that could affect the results to be similar across the full design. We aimed at minimizing the influence of site-related confounding factors and therefore based our review on common garden, retrospective stand and retrospective single tree designs. The use of retrospective single-tree plots is restricted by the fact that trees in natural mixed stands are not randomly distributed in terms of soil properties, they are not always of similar age, and the influence of surrounding trees cannot be ruled out (Rothe et al., 2002). Paired adjacent stands offer better control for edge effects than single-tree plots, but differences in stand history and soil conditions are less controlled than in common gardens. Common garden experiments are preferable for studying tree species effects on SOC as it is possible to control the influence of stand age and siterelated factors to a great extent (Binkley, 1995; Vesterdal et al., 2008). The studies by Ovington (1954, 1956) of ‘‘development of woodland conditions under different tree species’’ in the UK are among the first examples of SOC assessments based on common garden experiments and they possibly served as inspiration for forest researchers to establish similar experiments in other countries. However, common garden experiments are rare and often relatively young; thus studies of tree species influence have also been conducted under canopies of single trees in old forests or in older paired stands where trade-offs may be differences in site or soil type (Neirynck et al., 2000; Guckland et al., 2009). Common garden designs may also have a legacy of disturbances associated with cultivation which could bias understory plant composition as well as soil organisms in a direction that is not fully representative of the communities found in a natural setting. As a result of these pros and cons in terms of practicalities and statistical design, all three designs have been widely used. They are, however, all preferable to inventory designs such as in regional or national forest inventories in which tree species effects are generally strongly confounded with site-related factors that also potentially control SOC stocks.

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3. Is there consistent evidence of specific tree species effects from different experimental designs? We reviewed a number of studies from common garden, single tree, and paired stand studies to explore whether C stocks consistently differ in forest floors and top mineral soil under different tree species. Table 1 summarizes a number of SOC studies within similar genera of tree species in temperate and boreal forests. Of the studies that assessed forest floor C stocks all but one reported significant differences among tree species. The only exception was a quite young common garden experiment (aged 27 years) in the boreal region (Alriksson and Eriksson, 1998). It was remarkable that species rankings reported for forest floor C were relatively consistent, even between European and North American studies of the same broadleaf genera. In the temperate region, Fagus accumulated more C in forest floors than Fraxinus, Acer, and Tilia, whereas Quercus sp. was intermediate. In all cases, conifers (Picea

or Tsuga and in one case Larix) had highest C stocks in forest floors or were at least similar to Fagus. In the case of dominant boreal forest species, the pattern was also quite consistent. Forest floor C stocks were lower for broadleaves (Betula and Populus) than for Picea, while Pinus was intermediate but still close to Picea in C stock. A few studies in both temperate and boreal regions indicated that the deciduous conifer Larix did not consistently differ from either conifers or broadleaves in forest floor C stock (Table 1). Fewer studies focused on mineral soil C but species effects on this C pool appear weaker and less consistent. In the temperate region several studies found no significant influence, but in some cases there was a tendency of more C under certain hardwood tree species, i.e. Fraxinus, Acer, Tilia and Ulmus than under conifers and Fagus (Finzi et al., 1998a; Oostra et al., 2006; Vesterdal et al., 2008; Langenbruch et al., 2012). Thus, species with low C stocks in forest floors tended to have more C in mineral soil. However, Gurmesa et al. (this issue) found the same ranking of two broadleaf and

Table 1 Summary of SOC studies in similar genera of tree species in temperate and boreal regions of Europe and North America. Tree species differences are based on reported statistical significance (P < 0.05). Design: CG = common garden, PS = paired stand, ST = single-tree. Forest floor C stock Temperate forests 1 Quercus < Fagus, Pseudotsuga < Larix, Piceaa 2 Quercus < Picea < Larix, Pinusb 3 Fraxinus, Acer < Fagus < Quercus, Tsuga 4

Quercus < Fagus < Picea < Pinus

5 6 7

Fraxinus, Tilia, Acer < Quercus, Fagusa,b Betula < Pseudotsuga < Pinus Fraxinus < Fagus, Tsuga Acer < Tsuga (Quercus in between) NA

8 9 10 11

Acer < Tilia, Fagus, Quercus, Betula, Larix < Piceaa,b Fraxinus, Ulmus, Carpinus < Quercus < Fagus < Picea Fraxinus, Acer, Tilia < Quercus < Fagus < Picea

12 13

Tilia, Quercus < Pine, Picea, Larix Fagus < Pinus < Picea, Pseudotsugab

14 15 16 17

Quercus < Pinus Fagus, Quercus < Pseudotsuga, Piceaa Betula < Pinus < Picea Fraxinus, Tilia < Fagus


Quercus < Fagus < Picea < Larix

Boreal forests 19 Populus < Pinus, Piceab 20 No difference (Betula, Pinus, Larix, Picea) 21 Betula, Larix < Pinusb 22 Betula < Piceab 23 Populus < Piceaf 24 NA 25 NA 26 Populus < Picea 27 Betula < Picea (Pinus in between)e a b c d e f

Mineral soil C stock (to max. 30 cm)




No clear difference



Ovington (1954, 1956)

Quercus, Larix < Pinus < Picea 0–7.5 cm: No difference 7.5–15 cm: Fagus < Fraxinus, Acer NAc



Son and Gower (1992) Finzi et al. (1998a)



No difference NA No difference

Belgium Canada USA


Vesterdal and RaulundRasmussen (1998) Neirynck et al. (2000) Thomas and Prescott (2000) Dijkstra and Fitzhugh (2003)

No difference (Fraxinus, Fagus, Betula, Tilia, Quercus, Picea) No difference

Sweden, Lithuania, Denmark Poland


Hagen-Thorn et al. (2004)


Indicative: Picea, Fagus, Carpinus < Quercus < Ulmus 0–15: No difference



Reich et al. (2005), Mueller et al. (in press) Oostra et al. (2006)



Vesterdal et al. (2008)

Czech Republic France


Frouz et al. (2009) Mareschal et al. (2010)

15–30 cm: Picea < Fraxinus, Tilia (others in between) Picea, Pinus, Quercus, Larix < Tilia 0–5 cm: No difference 5–10 cm: Picea < Fagusd 10–15 cm: No difference Quercus < Pinus NA No difference 0–10 cm: Fagus, Tilia < Fraxinus 10–20 cm: No difference Quercus, Fagus < Picea, Larix

Spain France Sweden Germany


Díaz-Pinés et al. (2011) Trum et al. (2011) Hansson et al. (2011) Langenbruch et al. (2012)



Gurmesa et al. (this issue)

Populus < Pinus, Picea No difference No difference NA NA Larix < Pinus < Picea No difference (Betula, Larix) Populus < Picea No difference

USA Sweden Iceland Finland Canada Iceland Iceland Canada Finland


Alban (1982) Alriksson and Eriksson (1998) Sigurðardóttir (2000) Smolander et al. (2005) Legaré et al. (2005) Bjarnadóttir (2009) Ritter (2007) Laganière et al. (2011, 2012) Olsson et al. (2012)

No statistics given. Based on forest floor mass. NA: not assessed. Carbon concentrations. Significant difference in total soil C mainly driven by forest floor C. Forest floor thickness.

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Proportional difference in forest floor C

two coniferous species in both forest floor and mineral soil C stocks, and Mueller et al. (in press) found a nearly significant (P = 0.07) effect of tree species on C concentration in the A horizon (highest under Picea abies, lowest under Pinus nigra and Tilia cordata), although soil sampling on a depth basis to 20 cm showed no significant effect on C stock (P > 0.1). Few studies have reported on mineral soil C stocks in the boreal region (Table 1), which makes conclusions difficult. The opposite trend for forest floor and mineral soil C stocks in temperate species comparisons was not evident in the boreal forest comparisons where the species ranking for forest floor C stock and mineral C stock was more consistent between genera, i.e. higher C stocks in forest floor as well as mineral soil under conifers (Picea, Pinus, Larix) than broadleaves (Betula, Populus). We explored the relative maximum magnitude of the tree species effect in the 27 studies reviewed in Table 1 by calculating the ratio between the tree species with the significantly highest and the tree species with the significantly lowest forest floor or mineral soil C stock (Fig. 1). Forest floor C stocks showed the greatest proportional difference; the majority of studies in temperate forests showed more than 2-fold differences between the species with least and most forest floor C, but 4–10-fold differences were found in 10 studies of 24 that measured forest floors. These 10 studies all included comparisons of Acer/Fraxinus/Quercus/Betula vs. coniferous species. The two most extreme studies with 45- and 70-fold

differences included species with virtually no forest floor material (Ulmus and Tilia) compared to Picea. On the other hand, five of six studies with <2-fold difference were from the boreal region. These studies only reported a 20–70% increase in forest floor C stock from Betula or Populus to Picea. In the mineral soil, proportional differences were generally much lower within the range 1.2–1.8 for all but one study conducted in a former mine spoil area with extremely low pre-planting C stocks. In 9 studies of the 22 that assessed mineral soil C stocks, the proportional difference was not different from 1. It is important to note that the proportional differences in forest floors and mineral soils are not always additive – different species represent the maximum and minimum values for forest floor and mineral soil C. Our figures for forest floor C stocks are in accordance with the similar analysis of global forest floor mass data by Binkley and Giardina (1998). They reported forest floor mass differences among species of around 20% but frequently up to 5-fold. We are not aware of any other analyses of proportional differences in the top mineral soil C pool which constitutes the greatest amount of SOC in boreal and temperate regions (Callesen et al., 2003). This analysis indicates that a species effect of up to ±50% in the top mineral layer is not unrealistic to expect from a change in tree species. It seems safe to conclude that tree species influence on mineral soil C stocks is weak and less consistent. However, this conclusion

70 40 10

6 5 4 3 2 1 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27

Study id

Proportional difference inmineral soil C







1 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27

Study id Fig. 1. Proportional differences between the maximum and minimum stock of C in (a) forest floors and (b) top mineral soils under different tree species based on studies listed in Table 1. Studies 1–18 are from temperate forests and studies 19–27 are from boreal forests. Missing values are due to either missing data ((a) 3 studies, (b) 4 studies) or insignificant (P > 0.05) effects of species ((a) 1 study, (b) 9 studies).

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must be tempered by the fact that there are few long-term multispecies balanced common garden designs with an age >40– 50 years. As management-related effects, including that of tree species selection, generally take much longer to develop in mineral soil than forest floors (Jandl et al., 2007; Mueller et al., in press), it may be premature to assess effects of individual tree species based on young stands in common garden experiments regardless of the virtues of such designs. In this respect, the single-tree studies enable assessment of soils subject to tree species influence for a longer time. However, single-tree studies suffer more from influence of surrounding trees, particularly in broadleaved forests where windtransportation of litterfall is more important (Rothe et al., 2002; Langenbruch et al., 2012), and location of trees species is not independent of soil conditions. In some single-tree studies in older forests mineral soil C stocks indeed differed between species (Finzi et al., 1998a; Langenbruch et al., 2012, Table 1), but in other studies there were also no effects on soil C stocks or organic matter content (Neirynck et al., 2000; Dijkstra and Fitzhugh, 2003; Nordén, 1994). The effect of species on mineral soil C may be stronger or more similar in direction to that of forest floor C in boreal forests, but conclusions are hampered by paucity of data. The paired stands of trembling aspen (Populus tremuloides Michx.) and black spruce (Picea mariana (Mill.) BSP) forests in Canada studied by Laganière et al. (2011, 2012) and the common garden study in Minnesota, USA by Alban (1982) are the few cases in which C stocks were reported significantly lower under Populus than Picea in top mineral soil. While aspen in Canada had 35 and 62 Mg C ha 1 in mineral soil and forest floor plus mineral soil, respectively, black spruce stands stored 46 and 86 Mg C ha 1, respectively. In Minnesota, Populus plots had 10–40% less C in 0–15 cm and 28–38% less C in 10–25 cm than Picea and Pinus plots (Alban, 1982). In Sweden mineral soil C (0–30 cm) decreased in the order Picea > Betula > Pinus, but differences were not significant (Hansson et al., 2011). Nitrogen-fixing tree species have been reported to be associated with higher soil C concentrations or stocks in a range of studies (Johnson, 1992; Johnson and Curtis, 2001). In temperate and boreal forests this trait is most commonly represented by alder (Alnus sp.) as well as black locust (Robinia pseudoacacia L.) but few multi-species common garden experiments or paired stand experiments with information on C stocks included these tree species (Table 1). As an exception, the study by Frouz et al. (2009) also included Alnus, but while this species had some of the highest mineral soil C contents it did not differ from the non-N-fixing deciduous species lime (T. cordata L.). Based on a meta-analysis, Johnson and Curtis (2001) reported that presence of N-fixing species (e.g. red alder (Alnus rubra Bong.) and black alder (Alnus glutinosa L.)) caused a marked positive change in the C concentration of mineral soil A horizon. Particularly for alder species, results from common garden experiments would be preferable to assess the effects on soil C since alder is often associated with moist or wetter soil types in natural stands, i.e. this could imply confounded effects of moisture and species in single-tree studies. Most evidence of higher soil C contents in temperate N-fixing tree species seems to come from comparisons of Douglas-fir (Pseudotsuga menziesii (Mirb.) Franco) and red alder in north-western USA and Canada. Cole et al. (1995) reported that red alder stands had 30% more C in forest floor and mineral soil than adjacent Douglas-fir stands after 50 years. This is consistent with the review by Johnson (1992) indicating that N-fixing species in North America with a few exceptions had 20–100% more C in soils than non-N-fixing tree species. In a paired stand comparison across the border between red alder and Douglas-fir stands in British Columbia, Lavery et al. (2004) found higher C concentrations in mineral soil under alder, but only in 45-yearold stands compared to no difference for 10- and 25-year-old stands. While the body of literature supports a difference in soil C between temperate alder and Douglas-fir stands, a similar effect


of N fixing species remains to be found and specifically related to the N-fixation trait for European temperate tree species.

4. Do tree species affect soil C distribution rather than C stock? Common garden assessments of total C stock (forest floor + mineral soil) are not abundant; in several cases forest floor and mineral soil C have not been reported concurrently (Table 1). As separation of forest floor and mineral soil is difficult to standardize and is subject to person-related bias among studies, combined assessment of mineral soil and forest floor C stocks is highly preferable for assessment of SOC differences. Recent studies have addressed soil C stocks in both forest floors and top mineral soil, thus providing a more complete assessment of tree species effects on soils. One of the few studies that included the widely distributed tree species European silver birch (Betula pendula Roth.), Norway spruce (P. abies (L.) Karst.) and Scots pine (Pinus sylvestris L.) found that the differences in forest floor C (birch < pine < spruce) were mainly responsible for forest floor plus topsoil (0–30 cm) C stocks (Hansson et al., 2011). In this case forest floor C accounted for significant amounts of total C stocks, i.e. 15%, 34% and 44% for birch, pine and spruce, respectively, and thus played a larger role for total C stocks than in many broadleaf species found in temperate regions. Similar consistent findings for forest floor and mineral soils were reported for boreal lodgepole pine (Pinus contorta Dougl.) compared to downy birch (Betula pubescens Ehrh.) and Siberian larch (Larix sibirica Ledeb.) in eastern Iceland (Sigurðardóttir, 2000). In this study the forest floor similarly contained 19%, 21% and 25% of total soil C for birch, larch and pine, respectively. In a few temperate forest studies, the ranking of species with regard to C stocks was similar in both forest floor and mineral soil (Díaz-Pinés et al., 2011; Gurmesa et al., this issue, Table 1), but this was not the general pattern. In contrast, several other studies indicated that patterns in forest floor C stock may not always reflect C stocks in the mineral soil profile (Table 1) and that the sum of forest floor C and mineral soil C stocks tends to be similar among species. In fact, intriguing results from common garden and single tree studies suggest that tree species that accumulated most C in forest floors tended to have less C in top mineral soil. In a Danish common garden experiment including six tree species (Vesterdal et al., 2008) there was a negative relationship between forest floor C and mineral soil C in 15–30 cm (Fig. 2a). The differences in forest floor C contents among tree species appeared to be offset by differences in mineral soil C which explained the similarity in total C stock among tree species. Along the same line, Frouz et al. (2009) found that the amount of C in the top mineral soil layer (A horizon) under six tree species planted on reclaimed mine spoil was negatively correlated with C stock in the forest floor. However, in this study the pools did not offset each other and certain broadleaf species (lime and black alder) had more total soil C than Norway spruce and Scots pine (Fig. 2b). Langenbruch et al. (2012) found the same opposite trend between forest floor C and top mineral soil C stocks in the case of lime, ash (Fraxinus excelsior L.) and beech (Fagus sylvatica L.) in Germany. The opposite allocation of C and N in soil was also noted in other studies listed in Table 1, e.g. by Finzi et al. (1998a) in a study of hemlock (Tsuga canadensis Carr.) and North American species of maple, ash, beech, and oak, where the least forest floor C but the most mineral soil C were found under maple and ash. In a Swedish unreplicated common garden experiment, oak (Quercus robur L.) and ash had more C in the mineral soil than Norway spruce and beech, while the pattern was reversed for the forest floor (Oostra et al., 2006). In a French common garden experiment (Mareschal et al., 2010), beech stands were similarly reported to have less forest floor C but more mineral soil C than

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a) P = 0.028, R2 = 0.74

Total forest floor C, Mgha-1

Norway spruce 10

Beech Oak


Lime Ash

1 10




Mineral soil C (15-30 cm), Mg ha-1


Fig. 2. Soil carbon distribution between forest floor and mineral soil. (a) Relationships between forest floor C content and mineral soil C content in the 15–30 cm layer in a Danish common garden experiment (Vesterdal et al., 2008). (b) Carbon stocks in mineral soil (top graph) and forest floor (bottom graph) in reclaimed mine spoil stands after 22–30 years (Frouz et al., 2009). (c) Mean thickness of forest floor and A horizons in bigleaf maple and conifer plots in British Columbia (Turk et al., 2008).

Norway spruce and Douglas-fir stands. In Canada, Turk et al. (2008) reported that bigleaf maple (Acer macrophyllum Pursh) forest floors were thinner but with similar C stocks to plots dominated by Douglas-fir and western hemlock (Tsuga heterophylla (Raf.) Sarg.) (Fig. 2c). However, mineral soil A horizons were thicker and tended to contain more C under maple with the result that total contents of C did not differ between maple and conifer plots. Similar results were reported for paired stands of bigleaf maple and Douglas-fir in Oregon by Fried et al. (1990). These results altogether suggest that distributions of C may differ substantially within common European and North American tree species and that inferential knowledge about total soil C stocks based on forest floor C stocks may prove invalid. The apparent differences in C distribution rather than total C stock in certain 30–40 year-old common garden experiments suggest we temper our expectations regarding the short- to mediumterm role of tree species selection for mitigation of atmospheric CO2. Prescott (2010) suggested that to better manage soil C stocks based on knowledge about litter decomposition focus should be drawn to the diversion of litter into stable pools rather than just trying to slow the speed of decomposition. The diversion of new tree species-derived C into mineral soil rather than forest floor and mineral soil is particularly striking in the reclaimed mine spoil study by Frouz et al. (2009), but several studies reviewed here indeed suggest that stability of C stocks toward disturbances may be manipulated. Carbon sequestered in the forest floor is more vulnerable to increased decomposition following disturbances, e.g. clearcutting or burning during forest fires (Jandl et al., 2007). Mineral soil C pools are better protected, particularly in clay-rich soils, than are forest floor C pools (von Lützow et al., 2006; Jandl et al., 2007), but even within the mineral soil C pool, tree species differences in C stability have been reported. In boreal trembling aspen-black spruce forest Laganière et al. (2011) found more SOC in the less protected fractions, i.e. uncomplexed organic matter, under spruce than under aspen, the mixed stands being intermediate. Mueller

et al. (in press) found a correlation between SOC and total extractable acidity in the mineral soil, suggesting that tree species might affect SOC content in the mineral soil through species-specific effects on acidification and thus mineral weathering. Thus selection of tree species that better divert litter C into stable mineral soilassociated humus and not only less protected forest floor C would be one way of managing for C sequestration. While certain broadleaf and conifer species in temperate forests seem to differ in this respect, we need more studies from boreal forests to support conclusions whether tree species here primarily differ in superficial forest floor C stocks or also differ in mineral soil C stocks as reported by Laganière et al. (2011, 2012). The frequently observed opposite patterns in forest floor and mineral soil C contents (Fig. 2) stress the necessity to assess both the mineral soil and the forest floor in order to evaluate dynamics in C allocation. Ignorance of mineral soils would often result in misinterpretation of soil C sequestration under different tree species. The current insufficient knowledge of tree species effects on mineral soil C may partly be attributed to inadequate attention to soil types and sampling scheme (Jandl et al., 2007).

5. Mechanisms and processes responsible for SOC stock differences and SOC distribution The specific mechanisms and processes behind tree species differences in soil C stocks, including the apparent trade-off in soil C stock between forest floor and mineral soil, mostly remain unexplored, but several mechanisms may be in play. Soil C stocks represent the balance between input of C by litter production and output of C by decomposition and associated heterotrophic respiration of CO2, as well as leaching of dissolved organic carbon (DOC). Input of C from plants occurs above- as well as belowground and various processes and organisms can divert C inputs to different layers of the soil.

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Forest floor C stocks are directly affected by litterfall C inputs, but most studies of litterfall C inputs in common garden or paired stands found no or little difference in litterfall C among different tree species, and any difference was not likely to explain the large range in forest floor C stocks (Binkley and Valentine, 1991; Vesterdal et al., 2008; Hansen et al., 2009; Frouz et al., 2009; Trum et al., 2011). In a synthesis of tree species effects, Augusto et al. (2002) concluded along the same line that the annual amount of litterfall of a mature temperate forest stand is only slightly influenced by the species of the overstory because the major influences are climate and stand management. The above-mentioned studies suggest that tree species differences in forest floor C stocks are controlled by output processes, i.e. decomposer activity, rather than litterfall C inputs. Exceptions to this have been reported when basal area and canopy cover differ widely among species. In paired pure and mixed Scots pine and Pyrenean oak (Quercus pyrenaica L.) stands in Spain, Díaz-Pinés et al. (2011) found the lowest C stocks in forest floor and top mineral soil (0–5 cm) in oak stands in which litterfall C inputs and basal area were also lowest. In Sweden, low forest floor C stocks in paired adjacent birch, spruce and pine stands also appeared associated with corresponding differences in basal area and aboveground litterfall C input (Hansson et al., 2011). Although the abundance of ground vegetation differs among tree species (e.g. Barbier et al., 2008) the C input from this compartment is usually not known. Since broadleaf species support more ground vegetation than most conifer species, the litter input from ground vegetation may not explain the pattern in C accumulation between the species groups, but the litter production and its quality may contribute to differences within broadleaf as well as coniferous species. In the mineral soil the input flux of new or partly decomposed forest floor material may be driven by differences in soil fauna activity. Large variation in soil fauna composition and in particular earthworm abundance between broadleaf and conifer species and even within broadleaf species have been documented from common garden experiments (Reich et al., 2005; De Schrijver et al., 2012) and paired stands (Gudleifsson and Sigurdsson, in press). A comprehensive study of earthworm abundance under different tree species in Belgian and Danish common garden plots clearly showed that earthworm abundance or biomass increased from spruce over beech and oak to maple (Acer pseudoplatanus L.), lime and ash (Schelfhout, 2010). These tree species differences were also observed in paired stands by Neirynck et al. (2000) and in common garden plots by Reich et al. (2005) and De Schrijver et al. (2012) and were attributed to influences of tree species on litterfall Ca concentrations, soil pH and moisture. Earthworm abundance, particularly that of burrowing endogeic species (Marhan and Scheu, 2006; Felten and Emmerling, 2009), is a likely explanation for the observed differences in distribution of soil C stocks between certain broadleaves due to higher inputs of forest floor material to top mineral soil. This effect is likely to materialize within decades since the topsoil (0–15 cm) in temperate climates may be completely turned over every 10–20 years by earthworms (Edwards, 2009). In reclaimed post-mining soils, Frouz et al. (2009) also found that the tree species effect on soil C stocks was correlated positively and significantly with earthworm density and the abundance of earthworm cast in topsoil, suggesting that bioturbation could play an important role in soil carbon storage (Fig. 3). A second pathway for C incorporation in mineral soil may be via DOC leaching fluxes. Higher input of DOC from forest floor to mineral soil under spruce and pine than birch was reported from Sweden by Fröberg et al. (2011). The differences in DOC flux were primarily attributed to lower forest floor C stocks in birch (Hansson et al., 2011), but mineral soil C stocks were similar in the three tree species. In a common garden experiment in Denmark, Andersen


Fig. 3. Relationships between total C stocks and (a) earthworm casts and (b) earthworm abundance under various tree species in an afforested former mine spoil area (Frouz et al., 2009).

et al. (2004) similarly found DOC inputs to the top mineral soil to be highest under forest floor-accumulating species such as Norway spruce and beech. In such cases DOC input fluxes would therefore not seem likely to explain the apparent trade-off between forest floor and top mineral soil C stocks. Few studies have addressed the DOC flux pathway using common garden experiments, and species-related trends in DOC fluxes do not seem strong at larger spatial scales. In a review of DOC concentrations in temperate forests, Michalzik et al. (2001) found no general difference in DOC concentrations and fluxes in forest floor leachates between coniferous and hardwood sites, and there was also no relationship with forest floor C stocks. A third mechanism related to C input would be differences in root input due to differences in root biomass and turnover. Roots inputs are difficult and laborious to quantify and have thus been neglected compared to aboveground litterfall C input. However, it is known from synthesis studies in temperate regions that root litter may contribute to the soil C pool with an amount of C that equals that in foliar litterfall (Vogt et al., 1986; Kleja et al., 2008), and that deciduous forests often have higher root biomass (Finér et al., 2007). In simulation studies, Rasse et al. (2005) found that root litter inputs exceeded those of aboveground litter during 66 and 90 years in Scots pine and beech stands, respectively (Fig. 4). Moreover, stability of root-derived C may be higher than for foliar litter-derived C in certain ecosystems. Crow et al. (2009) concluded that differences in root productivity was the mechanism most likely to increase formation of stable C in their broadleaf forest site whereas aboveground needle input seemed more important in the coniferous site. These estimates emphasize that root litter inputs cannot be neglected in the attempt to disentangle tree species effects on forest soil C stocks and that this should receive more attention in future research. In addition to the magnitudes of above- and belowground litter inputs, the placement of litter input is a crucial factor. As reviewed by Schmidt et al. (2011), preferential retention of root-derived carbon has been observed in temperate forests, for example where below-ground inputs, including fungal mycelia, make up a larger

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Fig. 4. Simulated production of root litter C input over the lifespan of a 90-year-old beech forest in Belgium (Rasse et al., 2005).

fraction of new carbon in SOM than do leaf litter inputs. In addition to many above-ground inputs being mineralized in the forest floor, root and mycorrhizal inputs have more opportunity for physicochemical interactions with soil particles (Rasse et al., 2005; Schmidt et al., 2011). Multi-species comparisons of root biomass and root C input to soils in common garden experiments are rare and this constitutes an evident knowledge gap for interpretation of tree species differences in SOC stocks and C distribution. In one of the few multi-species but unreplicated common garden studies of root biomass, Oostra et al. (2006) reported that root biomass (<5 mm) and its proportion in the forest floor and mineral soil varied between tree species with lower root biomass in 0–20 cm depth under Norway spruce and beech than under oak and ash. Withington et al. (2006) performed an extensive replicated common garden study of root biomass and life span in topsoil under as much as 11 tree species in Poland and reported higher root density and root production in broadleaves than in conifers. Similar to the findings of Oostra et al. (2006), broadleaf root density decreased in the order maple > oak > beech = lime, and Norway spruce had higher root density than Scots pine. In a Swedish study, Hansson et al. (this issue) concluded that fine root depth distribution followed the same pattern as the C distribution in birch, pine and spruce, suggesting a significant contribution of root litter to SOC. Forest floor C stocks were closely related to root biomass (spruce > pine > birch), consistent with evidence that root systems of Norway spruce have a preference for forest floors (Puhe, 2003). However, the current studies could not determine whether high root biomass actively contributes to development of thick forest floors or is just an associated characteristic. Recent studies suggest that tree species influences via root dynamics should be studied with due attention to the type of associated mycorrhiza, i.e. ectomycorrhizal (EM) or arbuscular mycorrhizal (AM) associations (Phillips and Fahey, 2006). The presence or absence of mycorrhizal colonization is important for root turnover (Withington et al., 2006), and the decomposition rate of EM roots has been reported as slower than that of non-EM roots (Langley et al., 2006; Goebel et al., 2011). Communities of mycorrhizal fungi indeed differed significantly among tree species in common garden experiments (Grayston and Prescott, 2005; Phillips and Fahey, 2006; Buée et al., 2011), thereby suggesting functional effects of tree species-specific mycorrhizal associations in terms of C cycling and SOC stocks. Particularly in forest floors the input of mycorrhizal tissue may comprise a significant proportion of the organic matter input (Langley et al., 2006). A shift from AM to EM tree species could therefore be hypothesized to result in C accumulation.

There does indeed seem to be some association between EM species (e.g. spruce, pine and beech) and high forest floor C stocks, whereas AM tree species such as maple and ash have very low forest floor C stocks (Table 1). However, the functional effects of soil fungal communities and particularly mycorrhizal communities on both forest floor and mineral soil are not well known. Prescott and Grayston (this issue) reviewed the current knowledge of tree species influences on mycorrhizae and concluded that mycorrhizospheres could be a key in which tree species influence microbial communities. Such mechanisms would very likely contribute to differences in stocks and stability of SOC under various tree species. Soil aggregation, which can be affected by root production and turnover, has been shown to differ between tree species in a study including Picea, two Pinus species, Larix and Quercus (Scott, 1998). However, differences in the size distribution of the aggregates correlated only weakly with C mineralization, so aggregate formation is unlikely to be an important factor explaining species effects on C stocks. Root information in different tree species is still scarce and difficult to synthesize due to differences in estimation methods (Finér et al., 2007). Nevertheless high root biomass in spruce forest floors (Puhe, 2003) may contribute to explaining the large forest floor C stock in spruce stands, and the higher root biomass in mineral soil under ash and maple than under beech, spruce and oak (Ponti et al., 2004; Oostra et al., 2006; Withington et al., 2006) is well associated with higher C allocation to mineral soils under ash in several studies (Table 1; Fig. 2a and b). While litter input remains less studied within specific tree species, output of C from soils by decomposition and associated heterotrophic respiration has been addressed increasingly in research since development of the litterbag technique in the 1960s (Prescott, 2010) and recent focus on CO2 fluxes from forests (Subke et al., 2006). A plethora of studies have addressed decomposition rates of primarily foliar litterfall under different tree species, but again controlled assessments of tree species effects in common garden trials are relatively scarce (Binkley, 1995). However, more results have recently been published from common garden trials on foliar litter decomposition rates or heterotrophic respiration flux in different tree species (Ladegaard-Pedersen et al., 2005; Hobbie et al., 2006; Vesterdal et al., 2012). Large-scale paired stand comparisons have also confirmed differences in heterotrophic respiration, such as higher heterotrophic respiration in pure aspen and mixed aspen-black spruce stands than in pure black spruce stands in Quebec (Laganière et al., 2012). The general pattern is that litter quality traits control differences among tree species; however, the specific proxy most closely associated with such differences differs among studies depending on the tree species assemblages and thus the range in studied litter quality parameters. This follows from the fact that the measured factor with the greatest range in a given study is likely to become the best predictor providing it is related to decay rate, regardless of whether it has the greatest influence on rate of decay (Prescott, 2005). The most common proxies identified from common garden studies have been foliar litter concentrations of Ca (Reich et al., 2005), N or C/N ratio (Vesterdal et al., 2008, 2012), lignin (Hobbie et al., 2006) or lignin/N ratio. However, Hobbie et al. (2006) and Laganière et al. (2012) also noted the climatic differences related to tree species-specific stand structure as an important factor. Fast decomposition of forest floors has also been attributed to the presence of earthworms in certain tree species (Bohlen et al., 2004; Hobbie et al., 2006; Edwards, 2009) and thereby appears to be a strong determining factor for the magnitude of forest floor C stocks. An important interactive influence of soil fauna was indicated by paired stand and common garden studies in Sweden and Finland (Hansson et al., 2011; Olsson et al., 2012) in which birch

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forest floors were much lower in soil C compared to those of spruce and pine forest floors at the site with presence of earthworms. A similar interactive effect of earthworm presence and tree species (birch–Norway spruce) on C mineralization was also noted in microcosms by Saetre (1998). However, it remains a challenge to incorporate the biological activities of soil fauna in decomposition processes (Prescott, 2010), including differences noted between tree species. Most studies of tree species effects focused on recent litter decay using litterbags, but studies with concurrent use of litter bags and forest floor fractional loss (mass balance based on litterfall integrating also later-stage decay in forest floors) found lower average rates for mass balance than for litterbag decay (Hobbie et al., 2006; Vesterdal et al., 2012). Whereas the results for beech, oak, maple and lime were well related in a Danish common garden experiment (Vesterdal et al., 2012), Hobbie et al. (2006) found no significant relationship between the two estimates of forest floor decomposition rate for 14 tree species in a Polish common garden experiment. The species ranking as well as absolute values in forest floor fractional loss mass balance were very consistent between Poland and Denmark for maple, oak, lime, beech and spruce, suggesting that general species patterns in output flux of forest floor C do exist. Mineral soil C mineralization rates were not commonly reported from common garden plots, but the species effect appears relatively smaller. Consistently higher rates of C mineralization were observed under ash, maple and oak than under Norway spruce (Vesterdal et al., 2012), whereas Olsson et al. (2012) found no difference in mineral soil C mineralization rates under spruce, birch and pine. Less or different response to specific tree species could be explained by different sources of C controlling C turnover in forest floor and mineral soil; the influence of foliar litter input and its quality would be high for forest floor C turnover while the influence of root litter input and DOC would be more important for estimated mineral soil C turnover rates (Vesterdal et al., 2012). A recent study of above- and belowground litter decomposition in a Polish common garden experiment (Hobbie et al., 2010) indicated that we cannot expect species effects on aboveground litterfall to be reinforced in decomposition of belowground litter input. Influential litter traits differed between foliar and root litter and Hobbie et al. (2010) concluded that focus solely on aboveground traits may obscure the key mechanisms by which plant species influence ecosystem processes. It may therefore be unlikely to find similar controls on C mineralization within forest floor and mineral soil under different tree species. Despite the consistent and large effect of N-fixing tree species on the soil C concentrations or stocks, the fundamental processes remain largely unexplained (Johnson and Curtis, 2001; Resh et al., 2002). Is the positive effect derived from greater C inputs or reduced C outputs? Based on stable isotope fractionation in an afforestation experiment after sugarcane, Resh et al. (2002) reported that both mechanisms were in play: about 55% of the greater SOC sequestration under the N-fixers resulted from greater retention of old SOC4, and 45% resulted from greater accretion of new SOC3. Moreover, soil N accretion explained 62% of the variability of the greater retention of old SOC4 under the N-fixers. The latter result is in line with recent research indicating stabilization of SOC in recalcitrant forms under N rich conditions (Berg and Matzner, 1997; Janssens et al., 2010; Hobbie et al., 2012). Common garden studies with concurrent assessment of aboveground litterfall have shown that decomposition rates were more influenced by tree species than were litterfall inputs (Hansen et al., 2009; Trum et al., 2011; Vesterdal et al., 2008, 2012). This indicates that SOC stocks (particularly in the forest floor) were influenced through decomposition (output) differences rather than via differences in aboveground litterfall input. However, the num-


ber of studies is still too limited to support general conclusions regarding the main mechanisms (input or output) for tree species to influence forest floor and soil C stocks. Particularly for mineral soils, the many pathways and factors controlling C stocks (DOC flux, root C inputs and decay, macrofauna-mediated C mineralization and pedoturbation) are challenges for our current ability to understand tree species effects on soil C stocks.

6. Do results from studies at larger spatial scale match common garden, paired stand and single-tree studies? Some forest inventory programmes at regional (Vila et al., 2004) or national level (Berg et al., 2009) included soil sampling and thus provide information on SOC stocks under various tree species. However, we expect that the tree species signal will be diluted by influence of site-related factors compared to specific designs established to address tree species effects. Stendahl et al. (2010) reported an inventory study of Norway spruce and Scots pine stands in the national forest inventory (NFI) of Sweden covering boreal as well as temperate climate. The SOC stock was larger (92 Mg ha 1) for spruce than pine (57 Mg ha 1) under all temperature conditions, but the difference between species diminished with increasing temperature. These results led to the conclusion that Norway spruce should be selected on sites where both species are considered suitable if the aim of forest management is to maximize the soil C stock. The selection of stands strove to limit the confounding factor soil fertility (all sites were suitable for both species); however complete elimination of the site fertility factor was not possible given the inventory design. These results at the country scale are quite well in line with the paired stand scale results reported by Hansson et al. (2011) from southern Sweden (Table 1). Within European temperate forests, some regional inventories addressed tree species effects on SOC stocks. In a regional scale soil inventory in the Italian Alps Rodeghiero et al. (2010) reported that forest floor C stocks increased in the order beech < silver fir (Abies alba Mill.) < European larch (Larix decidua Mill.) < Norway spruce < Austrian pine (P. nigra Arnold) and Scots pine, but C stock differences were only significant between beech (8.7 Mg C ha 1) and spruce (22.0 Mg C ha 1). There were no significant differences in mineral soil C. In the Netherlands, Schulp et al. (2008) performed a soil inventory in beech, Douglas-fir, Scots pine, oak and larch stands planted on similar soil type at the forest scale (4200 ha) and reported a higher forest floor C stock in conifer than in broadleaf stands. Forest floor C stocks varied between 11.1 Mg C ha 1 (beech) and 29.6 Mg C ha 1 (larch) and mineral soil C stocks varied between 53.3 Mg C ha 1 (beech) and 97.1 Mg C ha 1 (larch). Forest floor results were in line with those from targeted tree species experiments (Table 1). Higher mineral soil C stocks under conifers than broadleaves were, however, not found in common garden studies that included the genera Fraxinus, Tilia and Acer, but are in line with studies including only conifers, beech and oak, and with studies in Sweden, Canada and Iceland of spruce and pine vs. birch and poplar (Table 1). The influence of European tree genera on forest floor mass was addressed in a meta-analysis by Gärdenäs (1998). The 59 European stands were dominated by the data on common garden forest floor mass published by Ovington (1954) but also included single stand data in a gradient from north Sweden to Spain. In spite of the large site-related variability, tree genus was the best single explanatory factor (r2 = 0.34). Forest floor mass in spruce stands (41 Mg ha 1) was significantly higher than those in larch, Douglas-fir, oak and birch stands (4–11 Mg ha 1). Beech deviated from the other deciduous species, birch, oak, grey alder (Alnus incana (L.) Moench) and chestnut (Castanea sativa Mill.), with a mean value of 19 Mg ha 1). This study is well in line with more recent common garden results for conifers, beech and

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oak (e.g. Vesterdal and Raulund-Rasmussen, 1998; Trum et al., 2011; Table 1), and also confirmed that Abies and Pseudotsuga appear to be intermediate in forest floor C stocks. In a chronosequence study of transformation of Norway spruce and Scots pine stands to Douglas-fir or beech in Germany, Prietzel and Bachmann (2012) found significantly decreased forest floor C stocks in the Douglas-fir or beech stands at 18 sites 30–120 years after conversion. Mineral soil C stocks were not significantly affected, but forest floor and mineral soil (0–50 cm) together also decreased significantly in C stock. The total SOC stock difference increased with age since conversion and was larger for conversions to beech than to Douglas-fir. The forest floor C dynamics inferred from the chronosequences are well in line with results from common garden studies in Table 1, but mineral soil C dynamics do not agree with the trend from some temperate common gardens that conifers store more C in forest floors and less in mineral soil than broadleaves. In conclusion, forest floor C in inventories and larger surveys are quite consistent with results from more targeted tree species designs, but trends in mineral soil C stocks do not clearly correspond to those from experiments. There is good evidence that the effect of soil type overrides that of tree species at larger spatial scale (Baritz et al., 2010).

7. Does tree species diversity affect soil carbon? It has for some time been hypothesized that increasing tree species diversity at stand level would lead to higher productivity at a given site (Scherer-Lorenzen et al., 2007). Two mechanisms have been suggested to induce higher productivity in mixed species communities based on agricultural systems (Vandermeer, 1989), (i) facilitation, i.e. a species promoting the growth and survival of another, mostly by improvement of abiotic conditions; and (ii) niche complementarity, i.e. a better and less competitive use of ecosystem resources between species with distinct functional traits (e.g. complementary degree of shade-tolerance or distribution of root system for exploitation of soil nutrients). There are indications that niche partitioning and complementary resource use or facilitation do occur in certain mixed forest types where interspecific competition was less than intraspecific competition (Cannell et al., 1992; Pretzsch, 2005) but effects on productivity are not consistently positive (Cavard et al., 2010). Given that more diverse, mixed stands have higher productivity this can be hypothesized to affect SOC stocks positively via higher litter production aboveground (Matala et al., 2008; Hansen et al., 2009) and belowground (Meinen et al., 2009a; Lei et al., 2012). While Lei et al. (2012) and Meinen et al. (2009a) found evidence of increased root production and turnover with increasing species diversity, i.e. higher C input to maintain soil C stocks, Meinen et al. (2009b,c) found no support for higher root biomass in mixed forests vs. pure beech forests. Given a higher input of C to soils in species-rich forest stands, higher rates of decomposition might possibly offset this effect resulting in similar soil C stocks. A wealth of studies has addressed mixed-species effects on decomposition of foliar litter (Rothe and Binkley, 2001). There have been no reports of a consistent additive or synergistic effect of species diversity on C emissions from soil by decomposition (Prescott et al., 2000; Hättenschwiler et al., 2005; Jacob et al., 2009) and thus the question remains whether increasing tree species diversity per se will affect soil C stocks. There is evidence supporting a significant role of species diversity for ecosystem functioning in grassland ecosystems, but the influence of diversity per se on ecosystem services in forests is largely unknown (Scherer-Lorenzen et al., 2007). To our knowledge, only two studies have addressed forest SOC stocks as affected by

tree species diversity, at the small scale (Guckland et al., 2009) or inventory scale level (Vila et al., 2004), respectively. In mixed broadleaf forest in Germany, Guckland et al. (2009) reported a negative effect of species diversity on forest floor C, i.e. lower forest floor C stocks in three- and five-species mixtures than in beechdominated stands (Fig. 5a). However, these results were also attributed to lower proportional admixture of beech, i.e. a species identity effect was probably responsible rather than species diversity per se. In the mineral soil, Guckland et al. (2009) also found a positive effect of tree species diversity on soil C stock in 10– 30 cm (Fig. 5b). This would be in accordance with other reports from common garden and single-tree studies of more C in mineral soil in the presence of broadleaves other than beech and a changed distribution of C from forest floor to mineral soil (Vesterdal et al., 2008; Langenbruch et al., 2012). However, it was also found that the presence of broadleaves other than beech in this natural forest was confounded with soil texture difference, thereby emphasizing the challenge in using single-tree studies in natural forests. In contrast to Guckland et al. (2009), Vila et al. (2004) reported a quite strong positive effect of tree species diversity (1–5 species) on forest floor mass in a regional inventory in northeast Spain. Depending on the functional identity of the dominant species, forest floor mass did not differ with species diversity (deciduous forest), increased steadily with diversity (coniferous-dominated forest) or just increased from one to more species (sclerophyllous forest, Fig. 6). Although a range of site-related variables were included in the analysis and appeared of weaker influence than tree species and species diversity, such inventories do not enable isolation of the species or species diversity factor to the same extent as common garden experiments. For instance it remains a challenge that the most species-diverse forests could also be more likely to occur on richer soils. In this case, Vila et al. (2004) did not detect a species diversity effect on productivity, but productivity had a weakly positive separate effect on forest floor mass. More studies of tree species diversity effects on ecosystem services including soil C stocks are expected in the near future (Scherer-Lorenzen et al., 2007). Based on German studies it appears that effects of species diversity per se are smaller than differences between individual tree species in beech-dominated forest types, and it is challenging to disentangle effects of species diversity in natural forest conditions. On the other hand, a strong species diversity signal was detected at the regional level in Spain through a multitude of other site factors. Proper assessment of species diversity effects per se requires new experimental platforms to avoid influence of site-related confounding factors. Such platforms should supplement existing designs and strongly support the work to disentangle pure species as well as species diversity effects on SOC.

8. Where to go: challenges in disentangling mechanisms behind C stock differences This synthesis has revealed that tree species indeed influence SOC stocks with quite consistent patterns in forest floor and less consistent patterns in mineral soil. As existing 30–40-year-old common garden studies become older and allow for resampling, we may hope to detect more clear patterns. However, many common garden plots have been designed too small for good assessment of pure species effects through a full rotation, e.g. 10  10 m (Challinor, 1968) or 20  20 m (Reich et al., 2005). As effects of neighboring stands (litterfall, shadow, etc.) increase substantially as stands mature this may limit the use of such experiments with time. Another challenge is the difference in sampling methodologies among studies in terms of sampling depth and separation of forest floor and mineral soil. Moreover, the var-

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Fig. 5. Tree species diversity effects on SOC. (a) Forest floor and mineral soil C stocks of German stands with different diversity levels (DL1 = 1 species (beech), DL2 = 3 species (DL1 + ash + lime), DL 3 = 5 species (DL2 + hornbeam + maple). (b) Forest floor C vs. beech abundance (Guckland et al., 2009).

Fig. 6. Forest floor mass as affected by tree species diversity (1–5 species) in stands dominated by different tree species categories in Catalonia, NE Spain (Vila et al., 2004).

iability in soil conditions and its co-varying effect (Vesterdal and Raulund-Rasmussen, 1998) necessitates a large number of proper replicated stands to assess statistically significant tree species effects along gradients in soil fertility and other abiotic factors. This is a challenge to ensure within each study site or even within countries. Therefore, joint cooperation, e.g. through a network of common gardens and paired sites using harmonized sampling and analyses would support even more general conclusions regarding the SOC stock signal of tree species. The most prominent challenge in understanding species effects on SOC is to disentangle the mechanisms behind observed SOC stock differences. We need to move from empirical studies to characterization of processes responsible for SOC stocks under different tree species. The key processes in focus would be input rates of C with plant litter and output rates of C by heterotrophic respiration. Output by leaching of DOC is probably quantitatively less important than heterotrophic respiration (Fröberg et al., 2011; Olsson et al., 2012). Input of C with aboveground litter tends to be comparable among species within similar climatic conditions with some exceptions, but root litter input may be more quantitatively important than aboveground litterfall (Rasse et al., 2005) and also appears to differ between deciduous and coniferous tree species (Oostra et al., 2006; Finér et al., 2007). Species differences in litter contribution from ground vegetation above- as well as belowground are probably of less importance but largely unknown. The information on tree species specific root biomass and turnover is yet disproportionate to that on aboveground litterfall, but hopefully rhizotron technology (Withington et al., 2006) or ingrowth core

techniques (Lei et al., 2012) can be developed to provide less time consuming and costly assessments of this source for SOC. Particularly, near-infrared reflectance spectroscopy or plant wax markers for prediction of species composition in root mixtures (Roumet et al., 2006; Lei and Bauhus, 2010) seems promising for facilitated assessment of root proportions in tree species diversity experiments and consequently for belowground processes in control of SOC stocks. It is uncertain whether aboveground or belowground litter input contributes the most to SOC stocks. The relative contribution of above vs. belowground plant material can help to identify the mechanisms behind tree species-driven SOC changes in mineral topsoils. The aliphatic compounds cutin and suberin are among the most recalcitrant plant structures in soil (Rasse et al., 2005) and have been proposed as biomarkers of above- and belowground biomass tissues, respectively (Mendez-Millan et al., 2011). Another possibility would be use of isotopic labelled aboveground litter (Kammer et al., 2012) to separate input fluxes between aboveand belowground litter sources. Such methods may also contribute to explain the transfer of forest floor C to mineral soil under different tree species by different mechanisms (Fig. 2). Kammer et al. (2012) found that around 50% of the annual C loss from the forest floor in a beech forest could be attributed to soil fauna-mediated transfer of new litter to the mineral soil while only 4–5% was transferred to the top mineral soil as DOC. Output of C has been studied using proxies such as litterbags or total soil respiration which do not integrate the heterotrophic output processes in the entire soil profile or separate between heterotrophic and autotrophic fluxes (Subke et al., 2006), respectively. Moreover decomposition of belowground litter is scarcely studied. It has been assumed that belowground litter quality mirrored that of aboveground litterfall, but recent findings that neither quality, tissue life span nor decomposition rates were consistent with aboveground patterns prevent inferences based on foliar litter quality (Withington et al., 2006; Hobbie et al., 2010). A dedicated research effort on root dynamics within different tree species seems justified to resolve the role of roots for SOC stock development under different tree species. The use of more integrated measures of SOC effluxes, i.e. measures that integrate heterotrophic activity in forest floor as well as top mineral soil would probably better gauge the patterns in SOC output processes among tree species. Recent efforts to partition total soil respiration into autotrophic and heterotrophic respiration are promising (Subke et al., 2006; Laganière et al., 2012), but have not yet been used in common garden experiments. A cross-cutting issue for input and output processes is the association of tree species with soil fauna. What is the role of associated soil fauna for development of SOC stocks via (i) placement of C within the soil profile and (ii) C efflux by decomposition? Studies

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of forest floor decomposition rates and C stocks have led to the belief that soil fauna, in particular earthworms, would decrease overall SOC stocks due to increased rates of decomposition. Recent research focusing on the forest floor as well as mineral soil has proved this perception too simplistic (Frouz et al., 2009). Consequently we should focus on the association between certain tree species and soil fauna that facilitates incorporation of SOC into mineral soil. We cannot expect these tree species-soil fauna associations to be similar among soil types as certain soil conditions are required for e.g. earthworms to be present (Edwards, 2009). Moreover, the soil fauna-mediated effect of tree species on SOC stocks and SOC distribution should ideally be addressed in experimental designs offering analysis of tree species as well as soil type effects. Even less known than the macrofauna community effect is the effects of tree species on SOC through their differences in soil fungal communities. In their review, Prescott and Grayston (this issue) emphasized the challenge to determine the extent to which differences in exudates from mycorrhizal roots of different tree species affect the composition of the microbial community in the rhizosphere and mycorrhizosphere, and also potentially in the bulk soil and forest floor. Another pending cross-cutting issue is the stability of the stored C. We need to account for the fact that some tree species that store more C may do so in less stable form or vice versa. Does it make sense to focus on SOC stocks alone in tree species evaluations or should we pay more attention to the diversion of SOC into stable pools (Prescott, 2010)? From the perspective of climate change mitigation by C sequestration, tree species facilitating diversion of more C into mineral soils would be preferable to species facilitating forest floor C sequestration as the latter pool is more susceptible to management changes and disturbances (Jandl et al., 2007). However, there is limited knowledge about the stability of stored C in mineral soils; is it in particulate form, complexed or in association with clay minerals under various tree species? The allocation of C inputs to various pools would be determined by pedogenic processes, possibly related to tree species-specific soil acidity (Mueller et al., in press). Again, the strong influence of site factors such as soil type and soil fauna would best be studied in common garden designs. Well-documented retrospective stand or singletree designs may also be useful, particularly when it can be verified that key parent material properties are similar under various species (Finzi et al., 1998b). The long-term perspective for establishment of common garden experiments is a challenge when new forest management issues emerge. The recent focus on strategies to adapt European forests to climate change (Kolström et al., 2011) calls for better knowledge of functional significance of species composition as well as species diversity, but we are short of suitable experimental platforms to disentangle the diversity signal from confounding factors such as environmental heterogeneity or species identity (Nadrowski et al., 2010). A large part of existing diversity effect studies on SOC in beech forest ecosystems used a complete dilution design in which it was impossible to separate true diversity effects from the effect of a particular species diluting the dominance of the matrix species (Guckland et al., 2009; Nadrowski et al., 2010). New research platforms, e.g. common garden experiments and retrospective stands, are therefore required to address effects of tree species diversity on SOC. For targeted use of tree species to sequester SOC, we still need more generalizable quantitative information on long-term SOC stock differences among different tree species in boreal and temperate forests. This includes studies to better determine the consistency of tree species effects along gradients in climate, soil type and other site-related factors. Such long-term estimates will require repeated assessment of SOC in tree species experiments as

they mature, possibly supplemented by process-based modelling studies. The controlled common garden ecosystems may provide a good picture of expected relative differences in SOC driven by management decisions on tree species change. However, to address the influence of climate change-driven tree species dynamics on SOC in more natural forest types we will most probably need to reconcile information from all three experimental designs, i.e. common gardens as well as retrospective paired stands and single-tree studies. Management decisions regarding species selection are based on a wide range of criteria apart from C sequestration, let alone C sequestration in the soil. The next step should be to evaluate SOC sequestration after a change in tree species as one of several ecosystem services (Weslien et al., 2009) to address possible synergies or trade-offs between SOC sequestration and other ecosystem services. 9. Conclusions We conclude that there is indeed evidence of consistent tree species effects on SOC stocks based on three common experimental approaches, i.e. common gardens, retrospective paired stands, and retrospective single tree studies. Effects were clearest for forest floor C stocks (23 of 24 studies) with consistent differences for genera common to European and North American temperate forests: Fagus, Quercus, Fraxinus, Acer, Tilia and conifers (Picea, Pinus or Tsuga), and for the boreal forest genera Picea, Pinus, Larix, Populus and Betula. Support for generalization of tree species effects on mineral soil C stocks was more limited, but significant effects were found in 13 of 22 studies that measured mineral soil C. Less consistency can be attributed to the paucity of multi-species balanced common garden designs with an age >40–50 years, but different soil sampling depth is another factor complicating syntheses of mineral soil C stocks. Proportional differences in forest floor and mineral soil C stocks among tree species suggested that C stocks can be increased by 200–500% in forest floors and by 40–50% in top mineral soil by tree species change. However, these proportional differences in forest floors and mineral soils are not always additive: C distribution between forest floor and mineral soil rather than total C stock tends to differ among tree species within temperate forests. This suggests that some species may be better engineers for sequestration of C in stable form in the mineral soil, but it is unclear whether the key mechanism is root litter input or macrofauna activity. Tree species effects on SOC in targeted experiments were most consistent with results from larger-scale studies for forest floor C pools whereas mineral soil C stocks appear to be stronger influenced by soil type or climate than by tree species at regional or national scales. There were indications that higher diversity could lead to higher C stocks, but the role of tree species diversity per se vs. species identity effects needs to be disentangled in rigorous experimental designs. For targeted use of tree species to sequester soil C, we must identify the processes related to C input and output, particularly belowground, that control SOC stock differences. We should also study forms and stability of C along with bulk C stocks to assess if certain broadleaves store C in more stable form. Joint cooperation is needed to support syntheses and process-oriented work on tree species and SOC, e.g. through an international network of common garden experiments. Acknowledgements This study received financial support from the Villum Foundation (L.V., P.G.). The work also benefited from the Nordic Network ‘Forest Soil C-sink’ funded by NordForsk.

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