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Belowground Carbon Cycling at Aspen FACE
Chapter 10
Belowground Carbon Cycling at Aspen FACE: Dynamic Responses to CO2 and O3 in Developing Forests Kurt S. Pregitzer1 and Alan F. Talhelm College of Natural Resources, University of Idaho, Moscow, Idaho, USA 1 Corresponding author: e-mail: [email protected]
Chapter Outline 10.1. Introduction 10.2. The Aspen FACE Experiment 10.2.1. Net Primary Productivity and Community Composition 10.2.2. Fine Roots Dynamics 10.2.3. Mycorrhizal Fungi
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10.2.4. Soil Respiration 10.2.5. Soil Organic Matter 10.2.6. Soil Microorganisms and Extracellular Enzymes 10.3. Conclusions and Implications Acknowledgements References
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10.1 INTRODUCTION Belowground processes such as root growth, soil respiration, and soil carbon (C) formation are fundamental components of forest C cycling. Soils are often the largest pool of C within forests (Pregitzer and Euskirchen, 2004) and soil respiration is the second largest terrestrial C flux (Bond-Lamberty and Thomson, 2010). On average, more than one-third of forest gross primary productivity is allocated belowground in order to sustain processes such as root respiration and the production of fine roots and mycorrhizae (Litton et al., 2007). However, although the negative impacts of tropospheric ozone (O3) Developments in Environmental Science, Vol. 13. http://dx.doi.org/10.1016/B978-0-08-098349-3.00010-4 © 2013 Elsevier Ltd. All rights reserved.
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exposure on plant growth and the potential for rising atmospheric carbon dioxide (CO2) concentrations to increase plant productivity have been studied for decades (Dunn, 1959; Rich, 1964; Siegenthaler and Oeschger, 1978), the effects of these gases on belowground processes such as root growth, soil C storage, and mycorrhizal productivity remain poorly resolved (Anderson, 2003; Hungate et al., 2009; Nikolova et al., 2010; Norby and Zak, 2011; King et al., 2013, this vol.; Kraigher et al., 2013, this vol.; Matyssek et al., 2013, this vol.). In part, this is because many belowground processes remain at the frontier of scientific research (Brantley et al., 2011) and some fundamental properties of belowground systems, such as the turnover rate of fine roots (Gaudinski et al., 2010) and the composition of soil organic matter (Schmidt et al., 2011), are still actively debated. Elevated CO2 and elevated O3 are each thought to affect belowground processes indirectly, with direct effects on leaf-level processes creating changes in plant physiology and productivity and shifting in belowground C inputs (e.g. Anderson, 2003; Zak et al., 2000; King et al., 2013, this vol.). This cascade
P
O
2. Leaf physiology (photosynthesis)
P
O
3. Aboveground biomass P
1. Atmosphere
O
4. Fine roots
P
Ambient
O
5. Mycorrhizae
Elevated CO2 Elevated O3 P
O
P
Predicted
O
Observed
9. Mineralization (respiration)
P
O
6. Litter production P P
O
8. Microbial biomass
O
7. Soil organic matter
FIGURE 10.1 Conceptual diagram of primary belowground processes. Bar graphs are qualitative and illustrate differences between ambient and the main effects of CO2 and O3. Arrows represent directional connections among pools and fluxes, with solid arrows representing predicted connections and the dashed arrows representing potential new connections based on the results from Aspen FACE.
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of effects is shown in Figure 10.1 and begins with changes in atmospheric composition (1) that impact leaf-level physiological processes, particularly photosynthesis (2). These changes in photosynthesis often affect aboveground growth (3), which creates parallel impacts on fine roots (4) and mycorrhizae (5). Fine roots, mycorrhizae, and aboveground growth contribute to the flux of litter into the soil (6), both through mortality and exudation. Litter inputs form soil organic matter (7), which is the substrate for heterotrophic microorganisms (8). Fine roots, mycorrhizae, and soil heterotrophs contribute to the mineralization (9) of organic compounds, releasing CO2 through respiration and making nitrogen available for plant uptake. Many of the early studies on the effects of elevated CO2 or O3 on plant growth were conducted in greenhouses or growth chambers (Ceulemans and Mousseau, 1994; Karnosky et al., 2005) and consequently, belowground processes such as root growth were often constrained by soil volume (Ceulemans and Mousseau, 1994). Scaling results from tree-level chamber studies to the ecosystem is also difficult (Ceulemans and Mousseau, 1994; Pye, 1988). For instance, there is wide inter-specific variation in growth in response to CO2 (Bazzaz et al., 1990; Tolley and Strain, 1984) and O3 (e.g. Dunn, 1959; Karnosky and Steiner, 1981; Tingey and Reinert, 1975), but the influence of this variation on ecosystem function is difficult to predict (Bradley and Pregitzer, 2007). The invention of Free-Air CO2 Enrichment (FACE) technology in the early 1990s (Hendrey et al., 1999) made it possible to expose whole forest stands to elevated concentrations of CO2, reducing limitations on experimental scale and allowing more accurate characterizations of soil–atmosphere linkages. However, insights into competitive interactions under elevated CO2 were limited because FACE technology was predominately applied to established forests, and particularly to single-species plantations (Norby and Zak, 2011). In addition, although FACE technology has been adapted to O3 research (Matyssek et al., 2010; Matyssek et al., 2013, this vol.), ecosystem-scale experiments quantifying the effects of O3 or the interactive effects of CO2 and O3 on forests are rare (Ainsworth et al., 2012; King et al., 2013, this vol.).
10.2 THE ASPEN FACE EXPERIMENT In this context, the Aspen FACE experiment was established in the northcentral United States (45.61 N, 89.51 W). This experiment was designed to test the influence of competitive interactions on the response of developing stands of trembling aspen (Populus tremuloides Michaux) to CO2 and/or O3 in a factorial randomized complete block design. This species of aspen is the most widely distributed tree in North America and common in several regions subject to high O3 exposures (Karnosky et al., 2003). At Aspen FACE, aspen was grown in either mixed species communities at equal densities with birch (Betula papyrifera Marsh.) or maple (Acer saccharum Marsh.)
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TABLE 10.1 Pre-Treatment (1997) Soil Physical and Chemical Properties Within the Top 10 cm of Mineral Soil Mean
SE
Sand (%)
55.6
0.9
Silt (%)
36.6
0.8
7.8
0.4
0.11
0.00
Soil texture
Clay (%) Gravimetric moisture content (y at water holding capacity) 2
Bulk density (mg m )
1.33
0.03
1
Total C (mg g )
15.30
0.84
1
1.19
0.07
Total N (mg g ) C:N 1
Extractable P (mg g ) 1
Cation exchange capacity (cmol(þ) kg ) Base saturation (%)
12.9
0.2
136.8
6.0
0.54 57.5
0.06 5.5
Values are means with standard error (SE).
or in mixed genotype communities (five clones varying in sensitivity to CO2 or O3), all representing common forest types in the region. All communities were planted with 1 m spacing in July 1997; fumigation began shortly thereafter and continued until mid-2009. During the 11 full growing seasons of this experiment, average fumigation concentrations were 532 ml l1 for CO2 (þ44% relative to ambient) and 46 nl l1 for O3 (þ30%; Zak et al., 2011) and the forest communities grew from small trees (<0.25 m tall) to dense stands (>8 m tall). Soils (Table 10.1) are Alfic Haplorthods with a sandy loam Ap horizon overlaying a sandy clay loam Bt horizon (Dickson et al., 2000). Here, we use the Aspen FACE experiment as a comprehensive case study to test the conceptual framework of how CO2 and O3 affect belowground C cycling. We briefly summarize the changes in aboveground productivity and then follow the path of C belowground, from fine roots and mycorrhizae to free-living soil microorganisms and soil respiration. We conclude by returning to the conceptual framework and putting the results into a broader scientific context.
10.2.1 Net Primary Productivity and Community Composition Over the first 3 years of the experiment, these forest communities responded to CO2 and O3 as predicted: elevated CO2 increased net primary productivity
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(NPP), while elevated O3 reduced NPP (King et al., 2005). These changes were consistent with leaf-level measurements of light-saturated photosynthesis, which were higher under elevated CO2 and lower under elevated O3 (Darbah et al., 2010). However, the effects of elevated O3 were smaller and less consistent than the effects of elevated CO2 (Kets et al., 2010). Interestingly, the negative effect of O3 on NPP began to diminish after 3–5 years (depending on the community; King et al., 2005) and during the final 3 years of the experiment, elevated O3 no longer reduced NPP (Zak et al., 2011). The recovery in NPP may be a consequence of compensatory growth of O3tolerant plants in response to the suppression of O3-sensitive species and genotypes (Zak et al., 2011). For instance, exposure to elevated O3 and ambient CO2 increased the relative importance of birch and maple over aspen (Kubiske et al., 2007). However, the effect of elevated O3 on leaf litter production (g m2) remained relatively constant over the last 7 years of the experiment, averaging 13% (Talhelm et al., 2012). In contrast, elevated CO2 had a sustained positive effect on NPP (þ25% in 2008; Zak et al., 2011) and leaf litter production (þ32% in 2008; Talhelm et al., 2012).
10.2.2 Fine Roots Dynamics In these developing forests, fine root (<1 mm in diameter) biomass increased quickly during the first half of the experiment before stabilizing over the last 4 years of the experiment (Figure 10.2). Within this overall pattern, the response of fine root biomass to the treatments was highly dynamic. In the
FIGURE 10.2 Fine root (<1 mm in diameter) biomass during the Aspen FACE experiment. Data from estimates in King et al. (2005), Pregitzer et al. (2006, 2008), and Zak et al. (2011) for the top 25 cm of mineral soil. Circles represent ambient O3, triangles represent elevated O3, empty symbols represent ambient CO2, and filled symbols represent elevated CO2. Error bars are 1 standard error.
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initial sampling (1999), elevated CO2 nearly doubled fine root biomass in the top 10 cm of mineral soil, but elevated O3 did not have a significant effect (King et al., 2001). In subsequent samplings in 2000 and 2002 (to 25 cm in depth; Figure 10.2), elevated CO2 created increases in fine root biomass that were smaller but consistent across all community types (averaging þ37%), while the response to elevated O3 was community-specific (King et al., 2005; Pregitzer et al., 2008; Figure 10.3). Elevated O3 decreased fine root biomass within the aspen-birch and aspen-maple communities in these two samplings, but O3 unexpectedly stimulated fine root biomass by more than 30% within the aspen-only community (King et al., 2005; Pregitzer et al., 2008). The response of fine root biomass to elevated CO2 was still positive in 2005 and 2008 (Zak et al., 2011), but this stimulation shrank further to 27% in 2005 and 11% in 2008. The effect of elevated O3 was positive across all communities in 2008 (þ4%), but the stimulation of fine root biomass in the aspen-only community was smaller (þ9%) than in previous estimates. Less is known about the response of roots deeper in the soil. However, Rhea and King (2012) sampled fine roots in the aspen-only and aspen-birch communities to 1 m depth in 2005 and found that neither treatment significantly affected fine root biomass below the top 30 cm of soil. It is interesting to contrast leaves and fine roots, the two most ephemeral major elements of plant biomass. Like roots, leaves initially displayed both rapid growth and large treatment effects, followed by a period of more subtle growth and smaller treatment effects (King et al., 2005; Talhelm et al., 2012). Leaf and fine root biomass both responded positively to elevated CO2 in all three communities and negatively to O3 in the aspen-birch and aspen-maple communities (Figure 10.2), but the positive response of fine roots to elevated O3 in the aspen community was not observed for leaves (Figure 10.2). Although this divergent response was unique among the three communities, the negative response to elevated O3 was consistently stronger for leaf biomass than for fine root biomass in all three communities (Figure 10.3). Similarly, the ratio of leaf area to fine root area had a positive, though less consistent, response to O3 when measured in 2005 (Rhea and King, 2012). The production and mortality of fine roots in the aspen community was estimated using minirhizotron observations from 2002 to 2005 (Pregitzer et al., 2008). Production rates (mm mm1) were not affected by CO2 or O3, but elevated O3 significantly increased root mortality rates in 1 year (2003). Consequently, changes in production and mortality biomass (g m2 year1) were mostly a function of changes in standing biomass (Pregitzer et al., 2008). These responses contrast with observations of leaf life span in the aspen-only community, in which leaf abscission was delayed by elevated CO2 and unaffected by O3 (Riikonen et al., 2008; Taylor et al., 2008).
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FIGURE 10.3 Relative treatment effect size (elevated/ambient) for leaf biomass and fine root (<1 mm in diameter) biomass for each community. Leaf biomass effects from King et al. (2005) and Talhelm et al. (2012); fine root data sources as in Figure 10.2. Triangles represent fine roots, circles represent leaves, empty symbols represent O3 effects, and filled symbols represent CO2 effects. Horizontal lines represent no treatment response.
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10.2.3 Mycorrhizal Fungi There are no estimates of the mycorrhizal biomass or hyphal turnover rates. However, insight into the influence of CO2 and O3 on mycorrhizae is available from observations of ectomycorrhizal sporocarps (2003–2006) within the aspen-only and aspen-birch communities (Andrew and Lilleskov, 2009) and analyses of fungal ribosomal DNA within the mineral soil in 2002 (Lesaulnier et al., 2008) and 2007 (Edwards and Zak, 2011). Elevated CO2 increased overall sporocarp production by an average of 96% in the aspenonly community and 70% in the aspen-birch community, principally by increasing the abundance of late-successional ectomycorrhizal taxa. Elevated CO2 also increased the species richness of these sporocarps. Notably, elevated CO2 increased the abundance of the early-successional ectomycorrhizal taxa Inocybe in the 2002 molecular analyses (Lesaulnier et al., 2008), but this change was not evident in 2007 (Edwards and Zak, 2011). Elevated O3 did not affect sporocarp species richness and only affected sporocarp production and community composition under ambient CO2 (Andrew and Lilleskov, 2009). Here, elevated O3 decreased sporocarp production by an average of 12% in the aspen-birch community and, in contrast to the response of fine root biomass, decreased sporocarp production by 48% in the aspen-only community. Elevated O3 affected sporocarp community composition under ambient CO2 by increasing the abundance of earlysuccessional taxa such as Inocybe and Paxillus. Assessments of fungal community composition in 2007 based on ribosomal DNA abundance provided more evidence that elevated O3 had increased the abundance of Inocybe, but in general, O3 effects on ectomycorrhizal community structure in these molecular analyses were small and independent of CO2 (Edwards and Zak, 2011). In addition to Inocybe, ectomycorrhizal communities within the soil under elevated O3 had relatively more species of Cortinarius and fewer Laccaria and Tomentella species (Edwards and Zak, 2011). Although Andrew and Lilleskov (2009) found more early-successional species under elevated O3 and more late-successional species under elevated CO2, there were clear differences in sporocarp community composition between the ontogenetically similar ambient and þCO2 þ O3 treatments.
10.2.4 Soil Respiration For the first 7 years of the experiment, soil respiration was measured only in the two most productive communities, aspen-birch and aspen-only. Elevated CO2 stimulated soil respiration in these two communities throughout the first seven growing seasons of the experiment (1998–2005), but the size of this stimulation varied by community, year, and season (King et al., 2001, 2004; Pregitzer et al., 2006). For instance, the annual CO2 effect during the first 4 years ranged from þ13% to þ41% within the aspen-only community and
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from þ43% to þ77% within the aspen-birch community (King et al., 2004). Elevated CO2 continued to have a smaller effect on soil respiration in the aspen-only community than the aspen-birch community throughout the first decade of the experiment (Pregitzer et al., 2006, 2008) and effect of elevated CO2 was not significant within the aspen-only community in 2002. Here, soil respiration rates were similar in the aspen-only and aspen-birch communities under elevated CO2, but soil respiration tended to be higher under ambient CO2 in the aspen-only community (Pregitzer et al., 2006, 2008). Later in the experiment (2005–2007), soil respiration was also quantified within the aspen-maple community. Patterns in growing season soil respiration in this community were similar to those in the aspen-birch community in terms of the size of the overall soil CO2 flux and the relative effects of elevated CO2. During these 3 years, the annual stimulation across all three communities of growing seasonal soil respiration by elevated CO2 ranged from 25% to 31% (Pregitzer et al., 2008). In general, the magnitude of the CO2 stimulation varied seasonally; it was strongest in the middle of the growing season when soil respiration was greatest and weaker during the spring and fall when soil temperatures were low (King et al., 2001; Pregitzer et al., 2006). Elevated O3 had much less consistent effects on soil respiration. There was no significant O3 effect on soil respiration in 1998, but in 1999 elevated O3 caused a 20% decrease in soil respiration late in the growing season (August to November; King et al., 2001). Elevated O3 significantly decreased soil respiration within the aspen-only community in 2002, but in the two subsequent years there was a significant CO2 O3 interaction (P < 0.01) in both communities and the effect of elevated O3 became positive under elevated CO2 (Pregitzer et al., 2006). Rates of soil respiration were consistently 5–10% greater in the þCO2 þ O3 treatment than in the þCO2 treatment from 2005 to 2007, but the CO2 O3 interaction was not significant because elevated O3 often only slightly decreased soil respiration under ambient CO2. This slight positive effect of O3 was most consistent in the aspen-only community, matching the response of fine root biomass (Pregitzer et al., 2008). In fact, soil respiration rates in 2005 were significantly correlated with fine root biomass across treatments and communities (r ¼ 0.72; P ¼ 0.008; Pregitzer et al., 2008). In addition to increasing the flux of CO2 from the soil, elevated CO2 also increased the partial pressure of CO2 within the soil (Karberg et al., 2005; King et al., 2001). The increase appears to have resulted primarily from greater rhizosphere C fluxes rather than changes in soil moisture or temperature that would reduce the diffusion of CO2 (Karberg et al., 2005; King et al., 2001). Additional CO2 within the soil is important because it can dissolve into the soil solution and increase the rate of two key processes: soil weathering and long-term C sequestration within oceanic carbonates (Kump et al., 2000). In measurements throughout the 2002 growing season, Karberg et al.
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(2005) observed that increases in the partial pressure of soil CO2 resulted in a 22% increase in dissolved inorganic carbon (DIC) and a 14% increase in the concentration of carbonic acid in soil solution. Soil solution alkalinity also increased by 210% (Karberg et al., 2005), likely due to an enhancement of primary mineral weathering by carbonic acid. Although elevated O3 did not affect either the soil partial pressure of CO2 or the concentration of DIC, we note that DIC concentrations in 2002 followed the same rank order as in soil respiration measurements made in subsequent years (2003–2004; Pregitzer et al., 2006): þCO2 þ O3 > þ CO2 > ambient > þ O3. In contrast to the changes in DIC, measurements of dissolved organic carbon during the first 2 years of the experiments found no clear effect of elevated CO2 or O3 (King et al., 2001).
10.2.5 Soil Organic Matter In addition to changing the amount of fine root and leaf litter, CO2 and O3 also influenced the chemical composition of plant litter and low molecularweight organic compounds in the surface mineral soil. In the mineral soil, exposure to elevated CO2 and ambient O3 created a marginally significant (P ¼ 0.08) increase in sugar concentrations (approximately þ10%) and had a larger effect on soluble phenolics (P ¼ 0.07, approximately þ40%), but there were no other treatment effects on sugar or soluble phenolic concentrations (Johnson and Pregitzer, 2007). These differences contrast with changes in plant litter chemistry. Concentrations of non-structural carbohydrates (sugars and starch) and structural carbohydrates (cellulose and hemi-cellulose) in fine roots collected in 2002 were not affected by the treatments (Chapman et al., 2005). Concentrations of sugars and soluble phenolics in leaf litter increased in response to CO2 (Liu et al., 2005; Parsons et al., 2008), but the greatest increases tended to be in the interaction (þCO2 þ O3) treatment rather than with exposure to elevated CO2 alone (þCO2). Soil amino acid concentrations responded individualistically to CO2 and to O3. Elevated CO2 increased aspartic acid concentrations, while elevated O3 decreased arginine and methionine concentrations. Elevated O3 decreased valine concentrations under ambient CO2, but had no effect under elevated CO2. The 11 other amino acids assayed were unaffected by the treatments (Johnson and Pregitzer, 2007). Averaged across all communities and treatments, soil C content within the top 20 cm of mineral soil increased from 3350 230 g m2 when first measured in 2001 to 4010 160 g m2 in 2008 (Talhelm et al., 2009). Although elevated CO2 consistently increased NPP (King et al., 2005; Zak et al., 2011), fine root production, and leaf litter inputs (Talhelm et al., 2012), there was no increase in soil C content (Figure 10.3). Soil C content was not significantly affected by elevated CO2 in the aspen-birch and aspen-maple communities. In the aspen-only community, soil C accrued more slowly under elevated CO2 than under ambient CO2 during the last several
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years of the experiment, creating a large and significant negative effect (1740 g m2) of elevated CO2 during the final year of the experiment (P ¼ 0.003; Figure 10.3). Likewise, although elevated O3 reduced NPP early in the experiment, it did not affect soil C content. The mass of the soil organic horizon (forest floor) was somewhat greater under elevated CO2 and smaller under elevated O3 in both 2004 and 2008, but these differences were not significant (Zak et al., 2011). The lack of soil C accumulation under elevated CO2 contrasts with studies of leaf litter and fine root decomposition (Chapman et al., 2005; Liu et al., 2009; Parsons et al., 2008), wherein elevated CO2 either slowed decomposition (Chapman et al., 2005; Parsons et al., 2008) or had no significant effect (Liu et al., 2009). For elevated O3, litter decomposition studies had mixed results: increased decomposition (Parsons et al., 2008), decreased decomposition (Liu et al., 2009; Parsons et al., 2008), or no effect (Chapman et al., 2005). Insight into the changes in C cycling within the mineral soil can be gained by isolating and quantifying different fractions of soil organic matter. Hofmockel et al. (2011) partitioned soil organic matter sampled from the top 20 cm of mineral soil in 2003, 2004, and 2007 into several physical fractions: coarse particulate (>250 mm), fine particulate (250–53 mm), and mineral-associated (<53 mm). The majority of soil C was found in the mineral-associated fraction, nearly 3000 g m2. The pools of coarse particulate and fine particulate organic matter were considerably smaller, with each accounting for approximately 500 g m2 of C. Although elevated CO2 did not have significant effects on any of these fractions overall, it did affect how the pools of coarse particulate and mineral-associated soil C changed through time. Coarse particulate C pools increased faster under elevated CO2 and were 19% greater under elevated CO2 in 2007. However, this increase was matched by a trend towards decreased mineral-associated C as the effect of elevated CO2 changed from þ7% in 2003 to 1% in 2007. Together, these trends suggest that while there is no significant overall effect on soil C content, the processing of soil C has been altered by elevated CO2 (Hofmockel et al., 2011). The effects of elevated CO2 on coarse particulate soil C also varied significantly by community, with a 6% decrease within the aspen-only community and a 23% increase in the aspen-birch community. There was a similar pattern in the pool of mineral-associated soil C, but these differences were not significant. Elevated O3 did not significantly affect the amount of C in these three soil fractions, but did increase the C/N ratio and decrease the pool of soil N within the coarse particulate and fine particulate fractions. The fossil-fuel derived CO2 used to create the elevated CO2 treatment is heavily depleted in 13C. This depletion is reflected in the d13C of plant tissue and creates an opportunity to distinguish ‘new’ C added to the soil since the beginning of the fumigation and ‘old’ pre-existing C. In the initial survey of these pools within the aspen-only and aspen-birch communities in 2001, the
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amount of total new soil C formed under elevated CO2 was 51% lower under elevated O3 (þCO2 þ O3 vs. þCO2; Loya et al., 2003). Using a chemical fractionation technique, Loya et al. (2003) also found that approximately onethird of this new C was recalcitrant acid-insoluble organic matter and that elevated O3 reduced the formation of new acid-insoluble C by 48%. When these soils were incubated, more C was respired in the interaction (þCO2 þ O3) treatment than under elevated CO2 alone, suggesting that the changes in soil C pools were a product of increased C losses rather than decreased C inputs (Loya et al., 2003). However, the negative effect of elevated O3 on soil C formation under elevated CO2 was considerably weaker within a wider analysis that included all three communities and measurements from 2001 to 2008 (Talhelm et al., 2009). Elevated O3 did not significantly decrease the amount of new soil C formed under elevated CO2 in the aspen-maple community in 2001 and the negative effects of O3 within the aspen-birch and aspen-maple communities weakened in subsequent measurements and were no longer significant. The input of new soil C under elevated CO2 was evident within each of the three physical soil C fractions, but inputs of new C to the fine particulate and mineral-associated soil C were offset by loss of old soil C (Hofmockel et al., 2011). Elevated O3 reduced the fraction of new C somewhat within each of the three fractions, but not significantly. However, elevated O3 slightly, but significantly, increased the amount of old C within the coarse particulate and fine particulate soil C pools (Hofmockel et al., 2011). Although this isotopic technique cannot be used to determine why elevated CO2 reduced total soil C content within the aspen-only community, it is notable that the largest loss of pre-existing ‘old’ soil C under elevated CO2 occurred within this community (Talhelm et al., 2009).
10.2.6 Soil Microorganisms and Extracellular Enzymes The lack of soil C accumulation under elevated CO2 was surprising, but consistent with observations of soil extracellular enzymes. Cellobiohydrolase and N-acetylglucosaminidase are extracellular enzymes important in the microbial breakdown of cellulose and chitin, respectively, which are two of the most abundant biomolecules within soil. Early in the experiment (1999), when the relative effects of elevated CO2 on fine roots and leaf litter were large, elevated CO2 stimulated the activity of cellobiohydrolase and N-acetylglucosaminidase by approximately 50% (Larson et al., 2002). A similar stimulation of these enzymes by elevated CO2 was also observed in 2001 and 2002 (Chung et al., 2006). These observations were consistent with laboratory incubations in which elevated CO2 increased the respiration of added 13C-labelled substrates for these enzymes (Phillips et al., 2002). Interestingly, the effect of elevated CO2 on cellobiohydrolase was considerably stronger under ambient O3 than under elevated O3 in both the field and the laboratory incubation, but this was not apparent for N-acetylglucosaminidase
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(Edwards and Zak, 2011; Phillips et al., 2002). The activities of other soil extracellular enzymes used to breakdown soil organic matter, including phosphatase, peroxidase, 1,4-a-glucosidase, 1,4-b-glucosidase, 1,4-b-xylosidase, and1,4-b-N-acetylglucose-aminidase also tended to be higher under elevated CO2 early in the experiment, but these differences were sometimes not significant (Chung et al., 2006; Larson et al., 2002). As the relative effects of elevated CO2 on fine roots and leaf litter declined, the stimulation of cellobiohydrolase and N-acetylglucosaminidase by elevated CO2 also declined and was no longer significant by 2008 (Edwards and Zak, 2011). The shrinking effect of elevated CO2 on N-acetylglucosaminidase resulted from an increase in enzyme activity under ambient CO2 rather than a decrease under elevated CO2, but the changes through time in cellobiohydrolase were less consistent (Edwards and Zak, 2011). Elevated O3 had a smaller impact on extracellular enzymes. Elevated O3 reduced cellobiohydrolase activity within the Oe horizon by 17% in 2008 (Edwards and Zak, 2011) and reduced the activity of 1,4-b-glucosidase by 25% in the surface mineral soil in 2001–2002 (Chung et al., 2006). Matching the inconsistent positive response of some enzymes to CO2, a similar group of enzymes tended to decrease under elevated O3. Elevated CO2 and elevated O3 each increased microbial (Larson et al., 2002) and fungal biomass (Chung et al., 2006), but not significantly. Although plant community type and soil horizon appear to be the dominant controls, elevated CO2 and O3 did influence microbial community composition within the mineral and organic soil (Chung et al., 2006; Edwards and Zak, 2011). For instance, elevated CO2 appears to increase the abundance of Sistotrema fungi, while elevated O3 shifted the composition of agaricomycete fungi (Edwards and Zak, 2011). Elevated CO2 also increased bacterial taxonomic richness (Dunbar et al., 2012; Lesaulnier et al., 2008), but decreased archaeal diversity and reduced the abundance of archaeal DNA by 50% (Lesaulnier et al., 2008). There is evidence that elevated CO2 altered the abundance of a number of individual bacterial and archaeal taxa (Dunbar et al., 2012; Lesaulnier et al., 2008). For instance, of 1313 individual bacterial taxa identified at Aspen FACE, 75 taxa significantly increased or decreased in abundance in response to elevated CO2 (Dunbar et al., 2012). However, the functional roles of individual taxa are poorly defined (Dunbar et al., 2012).
10.3 CONCLUSIONS AND IMPLICATIONS Belowground C cycling responded dynamically to the CO2 and O3 treatments within the developing forests at Aspen FACE. Often, these processes were consistent with changes in aboveground productivity (Figure 10.3). For instance, elevated CO2 increased photosynthesis and aboveground productivity, resulting in concomitant increases in fine root biomass, soil respiration, the partial pressure of soil CO2 and DIC. However, some responses diverged
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from expectations in several important ways, particularly for O3 (Figure 10.1). First, elevated O3 reduced leaf biomass more than fine root biomass in all communities, contrasting with reviews that have found decreased root:shoot ratios under elevated O3 (Anderson, 2003; Grantz et al., 2006; Wittig et al., 2009). Further, elevated O3 stimulated fine root biomass in the aspen-only community and stimulated soil respiration in all communities during the latter part of the experiment. A positive response of both fine root biomass and soil respiration was found in another free-air forest O3 experiment (Nikolova et al., 2010; Matyssek et al., 2013, this vol.) and, in general, the response of fine root biomass to O3 is highly variable in developing forests (King et al., this vol.). Likewise, mycorrhizal sporocarp production responded differently to CO2 or O3 than did fine roots, sometimes in the opposite direction. Finally, changes in productivity were not reflected in soil C content (Figure 10.4), as extracellular enzyme activity changed in parallel with litter inputs.
FIGURE 10.4 Relative treatment effect sizes for soil carbon content for each community. Empty symbols represent O3 effects and filled symbols represent CO2 effects. Significant (P < 0.05) CO2 effect denoted with asterisk.
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Overall, the differences between the predicted and observed effects on belowground processes illustrate the gaps in our knowledge of these processes and the value in long-term field experiments. Unfortunately, similar experiments (in scale and duration) are too few in number to know if the responses of belowground processes at Aspen FACE are universal or exceptional. Because of the large role of belowground processes in terrestrial C cycling (Bond-Lamberty and Thomson, 2010; Litton et al., 2007; Pregitzer and Euskirchen, 2004), accurate predictions of future terrestrial productivity and C sequestration will require further long-term experiments to achieve a more sophisticated understanding of how these processes respond to air pollution, climate change, and other global change factors. Without accurate predictions of future forest C cycling, the ability to design effective climate policy will be limited.
ACKNOWLEDGEMENTS The Aspen FACE experiment was supported by funding from the U.S. Department of Energy and by the USDA Forest Service. We sincerely thank Donald Zak, Dave Karnosky, Andrew Burton, Jud Isebrands, George Hendrey, Richard Dickson, John Nagy, and Keith Lewin for their invaluable contributions.
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Darbah, J.N.T., Kubiske, M.E., Nelson, N., Kets, K., Riikonen, J., et al., 2010. Will photosynthetic capacity of forest trees acclimate after long-term exposure to elevated CO2 and O3? Environ. Pollut. 158, 983–991. Dickson, R.E., Lewin, K.F., Isebrands, J.G., Coleman, M.D., Heilman, W.E., et al., 2000. Forest atmosphere carbon transfer and storages (FACTS-II): the aspen free-air CO2 and O3 enrichment (FACE) project: an overview. General Technical Report NC-214. USDA Forest Service North Central Research Station, St. Paul, MN, USA. Dunbar, J., Eichorst, S.A., Gallegos-Graves, L.V., Silva, S., Xie, G., et al., 2012. Common bacterial responses in six ecosystems exposed to 10 years of elevated atmospheric carbon dioxide. Environ. Microbiol. 14, 1145–1158. Dunn, D.B., 1959. Some effects of air pollution on Lupinus in the Los Angeles area. Ecology 40, 621–625. Edwards, I.P., Zak, D.R., 2011. Fungal community composition and function after long-term exposure of northern forests to elevated atmospheric CO2 and tropospheric O3. Glob. Chang. Biol. 17, 2184–2195. Gaudinski, J.B., Torn, M.S., Riley, W.J., Dawson, T.E., Joslin, J.D., Majdi, H., 2010. Measuring and modeling the spectrum of fine-root turnover times in three forests using isotopes, minirhizotrons, and the Radix model. Global Biogeochem. Cycles 24, GB3029. Grantz, D.A., Gunn, S., Vu, H.-B., 2006. O3 impacts on plant development: a meta-analysis of root/shoot allocation and growth. Plant Cell Environ. 29, 1193–1209. Hendrey, G.R., Ellsworth, D.S., Lewin, K.F., Nagy, J., 1999. A free-air enrichment system for exposing tall forest vegetation to elevated atmospheric CO2. Glob. Chang. Biol. 5, 293–309. Hofmockel, K.S., Zak, D.R., Moran, K.K., Jastrow, J.D., 2011. Changes in forest soil organic matter pools after a decade of elevated CO2 and O3. Soil Biol. Biochem. 43, 1518–1527. Hungate, B.A., Van Groenigen, K.J., Six, J., Jastrow, J.D., Luo, Y., et al., 2009. Assessing the effect of elevated carbon dioxide on soil carbon: a comparison of four meta-analyses. Glob. Chang. Biol. 15, 2020–2034. Johnson, R.M., Pregitzer, K.S., 2007. Concentration of sugars, phenolic acids, and amino acids in forest soils exposed to elevated atmospheric CO2 and O3. Soil Biol. Biochem. 39, 3159–3166. Karberg, N., Pregitzer, K.S., King, J.S., Friend, A.L., Wood, J.R., 2005. Soil carbon dioxide partial pressure and dissolved inorganic carbonate chemistry under elevated carbon dioxide and ozone. Oecologia 142, 296–306. Karnosky, D.F., Steiner, K.C., 1981. Provenance and family variation in response of Fraxinus americana and Fraxinus pennsylvanica to ozone and sulfur dioxide. Phytopathology 71, 804–807. Karnosky, D.F., Zak, D.R., Pregitzer, K.S., Awmack, C.S., Bockheim, J.G., et al., 2003. Tropospheric O3 moderates responses of temperate hardwood forests to elevated CO2: a synthesis of molecular to ecosystem results from the Aspen FACE project. Funct. Ecol. 17, 289–304. Karnosky, D.F., Pregitzer, K.S., Zak, D.R., Kubiske, M.E., Hendrey, G.R., Weinstein, D., Nosal, M., Percy, K.E., 2005. Scaling ozone responses of forest trees to the ecosystem level in a changing climate. Plant Cell Environ. 28, 965–981. Kets, K., Darbah, J.N.T., Sober, A., Riikonen, J., Sober, J., Karnosky, D.F., 2010. Diurnal changes in photosynthetic parameters of Populus tremuloides, modulated by elevated concentrations of CO2 and/or O3 and daily climatic variation. Environ. Pollut. 158, 1000–1007. King, J., Liu, L., Aspinwall, M., 2013. Forest ecosystem response to interacting elevated atmospheric CO2 and tropospheric O3: a synthesis of experimental evidence. Dev. Environ. Sci. 13, 179–208.
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King, J.S., Pregitzer, K.S., Zak, D.R., Sober, J., Isebrands, J.G., Dickson, R.E., Hendrey, G.R., Karnosky, D.F., 2001. Fine-root biomass and fluxes of soil carbon in young stands of paper birch and trembling aspen as affect by elevated atmospheric CO2 and tropospheric O3. Oecologia 128, 237–250. King, J.S., Hanson, P.J., Bernhardt, E., DeAngelis, P., Norby, R.J., Pregitzer, K.S., 2004. A multiyear synthesis of soil respiration responses to elevated atmospheric CO2 from four forest FACE experiments. Glob. Chang. Biol. 10, 1027–1042. King, J.S., Kubiske, M.E., Pregitzer, K.S., Hendrey, G.R., McDonald, E.P., Giardina, C.P., Quinn, V.S., Karnosky, D.F., 2005. Tropospheric O3 compromises net primary production in young stands of trembling aspen, paper birch and sugar maple in response to elevated atmospheric CO2. New Phytol. 168, 623–636. Kraigher, H., Bajc, M., Grebanc, T., Pritsch, K., 2013. Mycorrhizosphere complexity. Dev. Environ. Sci. 13, 151–177. Kubiske, M.E., Quinn, V.S., Marquardt, P.E., Karnosky, D.F., 2007. Effects of elevated atmospheric CO2 and/or O3 on intra-and interspecific competitive ability of aspen. Plant Biol. 9, 342–355. Kump, L.R., Brantley, S.L., Arthur, M.A., 2000. Chemical weathering, atmospheric CO2, and climate. Annu. Rev. Earth Planet. Sci. 28, 611–667. Larson, J.L., Zak, D.R., Sinsabaugh, R.L., 2002. Microbial activity beneath temperate trees growing under elevated CO2 and O3. Soil Sci. Soc. Am. J. 66, 1848–1856. Lesaulnier, C., Papamichail, D., McCorkle, S., Ollivier, B., Skiena, S., Taghavi, S., Zak, D.R., van der Lelie, D., 2008. Elevated atmospheric CO2 affects soil microbial diversity associated with trembling aspen. Environ. Microbiol. 10, 926–941. Litton, C.M., Raich, J.W., Ryan, M.G., 2007. Carbon allocation in forest ecosystems. Glob. Chang. Biol. 13, 2089–2109. Liu, L., King, J.S., Giardina, C.P., 2005. Effects of elevated atmospheric CO2 and tropospheric O3 on leaf litter production and chemistry in trembling aspen and paper birch ecosystems. Tree Physiol. 15, 1511–1522. Liu, L., King, J.S., Giardina, C.P., Booker, F.L., 2009. The influence of chemistry, production and community composition on leaf litter decomposition under elevated atmospheric CO2 and tropospheric O3 in a northern hardwood ecosystem. Ecosystems 12, 401–416. Loya, W.M., Pregitzer, K.S., Karberg, N.J., King, J.S., Giardina, C.P., 2003. Reduction of soil carbon formation by tropospheric ozone under increased carbon dioxide levels. Nature 425, 705–707. Matyssek, R., Wieser, G., Ceulemans, R., Rennenberg, H., Pretzsch, H., et al., 2010. Enhanced ozone strongly reduces carbon sink strength of adult beech (Fagus sylvatica)—resume from the free-air fumigation study at Kranzberg Forest. Environ. Pollut. 158, 2527–2532. Matyssek, R., Wieser, G., Fleischmann, F., Grunhage, L., 2013. Ozone Research, quo vadis? Lessons from the Kranzberg Forest Experiment. Dev. Environ. Sci. 13, 103–129. Nikolova, P.S., Andersen, C.P., Blaschke, H., Matyssek, R., Haberle, K.-H., 2010. Belowground effects of enhanced tropospheric ozone and drought in a beech/spruce forest (Fagus sylvatica L./Picea abies [L.] Karst). Environ. Pollut. 158, 1071–1078. Norby, R.J., Zak, D.R., 2011. Ecological lessons from free-air CO2 enrichment (FACE) experiments. Annu. Rev. Ecol. Evol. Syst. 42, 181–203. Parsons, W.F.J., Bockheim, J.G., Lindroth, R.L., 2008. Independent, interactive, and speciesspecific responses of leaf litter decomposition to elevated CO2 and O3 in a northern hardwood forest. Ecosystems 11, 505–519. Phillips, R.L., Zak, D.R., Holmes, W.E., White, D.C., 2002. Microbial community composition and function beneath temperate trees exposed to elevated atmospheric carbon dioxide and ozone. Oecologia 131 (2), 236–244.
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Pregitzer, K.S., Euskirchen, E.E., 2004. Carbon cycling and storage in world forests: biome patterns related to forest age. Glob. Chang. Biol. 10, 2052–2077. Pregitzer, K., Loya, W., Kubiske, M., Zak, D., 2006. Soil respiration in northern forests exposed to elevated carbon dioxide and ozone. Oecologia 148, 503–516. Pregitzer, K., Burton, A., King, J., Zak, D., 2008. Soil respiration, root biomass, and root turnover following long-term exposure of northern forests to elevated atmospheric CO2 and tropospheric O3. New Phytol. 180, 153–161. Pye, J.M., 1988. Impact of ozone on the growth and yield of trees: a review. J. Environ. Qual. 17, 347–360. Rhea, L.K., King, J.S., 2012. Depth-dependency of trembling aspen and paper birch small-root responses to eCO2 and eO3. Plant Soil 355, 215–229. Rich, S., 1964. Ozone damage to plants. Annu. Rev. Phytopathol. 2, 253–266. Riikonen, J., Kets, K., Darbah, J., Oksanen, E., Sober, A., et al., 2008. Carbon gain and bud physiology in Populus tremuloides and Betula papyrifera under long-term exposure to elevated concentrations of CO2 and O3. Tree Physiol. 28, 243–254. Schmidt, M.W., Torn, M.S., Abiven, S., Dittmar, T., Guggenberger, G., et al., 2011. Persistence of soil organic matter as an ecosystem property. Nature 478, 49–56. Siegenthaler, U., Oeschger, H., 1978. Predicting future atmospheric carbon dioxide levels. Science 199, 388–395. Talhelm, A.F., Pregitzer, K.S., Zak, D.R., 2009. Species-specific responses to atmospheric carbon dioxide and tropospheric ozone mediate changes in soil carbon. Ecol. Lett. 12, 1219–1228. Talhelm, A.F., Pregitzer, K.S., Giardina, C.P., 2012. Long-term leaf production response to elevated atmospheric carbon dioxide and tropospheric ozone. Ecosystems 15, 71–82. http:// dx.doi.org/10.1007/s10021-011-9493-z. Taylor, G., Tallis, M.J., Giardina, C.P., Percy, K.E., Miglietta, F., et al., 2008. Future atmospheric, CO2 leads to delayed autumnal senescence. Glob. Chang. Biol. 14, 264–275. Tingey, D.T., Reinert, R.A., 1975. The effect of ozone and sulphur dioxide singly and in combination on plant growth. Environ. Pollut. 9, 117–125. Tolley, L.C., Strain, B.R., 1984. Effects of CO2 enrichment and water stress on growth of Liquidambar styraciflua and Pinus taeda seedlings. Can. J. Bot. 62, 2135–2139. Wittig, V.E., Ainsworth, E.A., Naidu, S.L., Karnosky, D.F., Long, S.P., 2009. Quantifying the impact of current and future tropospheric ozone on tree biomass, growth, physiology and biochemistry: a quantitative meta-analysis of the ozone literature. Glob. Chang. Biol. 15, 396–424. Zak, D.R., Pregitzer, K.S., King, J.S., Holmes, W.E., 2000. Elevated atmospheric CO2, fine roots and the response of soil microorganisms: a review and hypothesis. New Phytol. 147, 201–222. Zak, D.R., Pregitzer, K.S., Kubiske, M.E., Burton, A.J., 2011. Forest productivity under elevated CO2 and O3: positive feedbacks to soil N cycling sustain decade-long net primary productivity enhancement. Ecol. Lett. 14, 1220–1226.
Belowground Carbon Cycling at Aspen FACE: Dynamic Responses to CO2 and O3 in Developing Forests Kurt S. Pregitzer1 and Alan F. Talhelm College of Natural Resources, University of Idaho, Moscow, Idaho, USA 1 Corresponding author: e-mail: [email protected]
Chapter Outline 10.1. Introduction 10.2. The Aspen FACE Experiment 10.2.1. Net Primary Productivity and Community Composition 10.2.2. Fine Roots Dynamics 10.2.3. Mycorrhizal Fungi
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10.2.4. Soil Respiration 10.2.5. Soil Organic Matter 10.2.6. Soil Microorganisms and Extracellular Enzymes 10.3. Conclusions and Implications Acknowledgements References
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10.1 INTRODUCTION Belowground processes such as root growth, soil respiration, and soil carbon (C) formation are fundamental components of forest C cycling. Soils are often the largest pool of C within forests (Pregitzer and Euskirchen, 2004) and soil respiration is the second largest terrestrial C flux (Bond-Lamberty and Thomson, 2010). On average, more than one-third of forest gross primary productivity is allocated belowground in order to sustain processes such as root respiration and the production of fine roots and mycorrhizae (Litton et al., 2007). However, although the negative impacts of tropospheric ozone (O3) Developments in Environmental Science, Vol. 13. http://dx.doi.org/10.1016/B978-0-08-098349-3.00010-4 © 2013 Elsevier Ltd. All rights reserved.
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exposure on plant growth and the potential for rising atmospheric carbon dioxide (CO2) concentrations to increase plant productivity have been studied for decades (Dunn, 1959; Rich, 1964; Siegenthaler and Oeschger, 1978), the effects of these gases on belowground processes such as root growth, soil C storage, and mycorrhizal productivity remain poorly resolved (Anderson, 2003; Hungate et al., 2009; Nikolova et al., 2010; Norby and Zak, 2011; King et al., 2013, this vol.; Kraigher et al., 2013, this vol.; Matyssek et al., 2013, this vol.). In part, this is because many belowground processes remain at the frontier of scientific research (Brantley et al., 2011) and some fundamental properties of belowground systems, such as the turnover rate of fine roots (Gaudinski et al., 2010) and the composition of soil organic matter (Schmidt et al., 2011), are still actively debated. Elevated CO2 and elevated O3 are each thought to affect belowground processes indirectly, with direct effects on leaf-level processes creating changes in plant physiology and productivity and shifting in belowground C inputs (e.g. Anderson, 2003; Zak et al., 2000; King et al., 2013, this vol.). This cascade
P
O
2. Leaf physiology (photosynthesis)
P
O
3. Aboveground biomass P
1. Atmosphere
O
4. Fine roots
P
Ambient
O
5. Mycorrhizae
Elevated CO2 Elevated O3 P
O
P
Predicted
O
Observed
9. Mineralization (respiration)
P
O
6. Litter production P P
O
8. Microbial biomass
O
7. Soil organic matter
FIGURE 10.1 Conceptual diagram of primary belowground processes. Bar graphs are qualitative and illustrate differences between ambient and the main effects of CO2 and O3. Arrows represent directional connections among pools and fluxes, with solid arrows representing predicted connections and the dashed arrows representing potential new connections based on the results from Aspen FACE.
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of effects is shown in Figure 10.1 and begins with changes in atmospheric composition (1) that impact leaf-level physiological processes, particularly photosynthesis (2). These changes in photosynthesis often affect aboveground growth (3), which creates parallel impacts on fine roots (4) and mycorrhizae (5). Fine roots, mycorrhizae, and aboveground growth contribute to the flux of litter into the soil (6), both through mortality and exudation. Litter inputs form soil organic matter (7), which is the substrate for heterotrophic microorganisms (8). Fine roots, mycorrhizae, and soil heterotrophs contribute to the mineralization (9) of organic compounds, releasing CO2 through respiration and making nitrogen available for plant uptake. Many of the early studies on the effects of elevated CO2 or O3 on plant growth were conducted in greenhouses or growth chambers (Ceulemans and Mousseau, 1994; Karnosky et al., 2005) and consequently, belowground processes such as root growth were often constrained by soil volume (Ceulemans and Mousseau, 1994). Scaling results from tree-level chamber studies to the ecosystem is also difficult (Ceulemans and Mousseau, 1994; Pye, 1988). For instance, there is wide inter-specific variation in growth in response to CO2 (Bazzaz et al., 1990; Tolley and Strain, 1984) and O3 (e.g. Dunn, 1959; Karnosky and Steiner, 1981; Tingey and Reinert, 1975), but the influence of this variation on ecosystem function is difficult to predict (Bradley and Pregitzer, 2007). The invention of Free-Air CO2 Enrichment (FACE) technology in the early 1990s (Hendrey et al., 1999) made it possible to expose whole forest stands to elevated concentrations of CO2, reducing limitations on experimental scale and allowing more accurate characterizations of soil–atmosphere linkages. However, insights into competitive interactions under elevated CO2 were limited because FACE technology was predominately applied to established forests, and particularly to single-species plantations (Norby and Zak, 2011). In addition, although FACE technology has been adapted to O3 research (Matyssek et al., 2010; Matyssek et al., 2013, this vol.), ecosystem-scale experiments quantifying the effects of O3 or the interactive effects of CO2 and O3 on forests are rare (Ainsworth et al., 2012; King et al., 2013, this vol.).
10.2 THE ASPEN FACE EXPERIMENT In this context, the Aspen FACE experiment was established in the northcentral United States (45.61 N, 89.51 W). This experiment was designed to test the influence of competitive interactions on the response of developing stands of trembling aspen (Populus tremuloides Michaux) to CO2 and/or O3 in a factorial randomized complete block design. This species of aspen is the most widely distributed tree in North America and common in several regions subject to high O3 exposures (Karnosky et al., 2003). At Aspen FACE, aspen was grown in either mixed species communities at equal densities with birch (Betula papyrifera Marsh.) or maple (Acer saccharum Marsh.)
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TABLE 10.1 Pre-Treatment (1997) Soil Physical and Chemical Properties Within the Top 10 cm of Mineral Soil Mean
SE
Sand (%)
55.6
0.9
Silt (%)
36.6
0.8
7.8
0.4
0.11
0.00
Soil texture
Clay (%) Gravimetric moisture content (y at water holding capacity) 2
Bulk density (mg m )
1.33
0.03
1
Total C (mg g )
15.30
0.84
1
1.19
0.07
Total N (mg g ) C:N 1
Extractable P (mg g ) 1
Cation exchange capacity (cmol(þ) kg ) Base saturation (%)
12.9
0.2
136.8
6.0
0.54 57.5
0.06 5.5
Values are means with standard error (SE).
or in mixed genotype communities (five clones varying in sensitivity to CO2 or O3), all representing common forest types in the region. All communities were planted with 1 m spacing in July 1997; fumigation began shortly thereafter and continued until mid-2009. During the 11 full growing seasons of this experiment, average fumigation concentrations were 532 ml l1 for CO2 (þ44% relative to ambient) and 46 nl l1 for O3 (þ30%; Zak et al., 2011) and the forest communities grew from small trees (<0.25 m tall) to dense stands (>8 m tall). Soils (Table 10.1) are Alfic Haplorthods with a sandy loam Ap horizon overlaying a sandy clay loam Bt horizon (Dickson et al., 2000). Here, we use the Aspen FACE experiment as a comprehensive case study to test the conceptual framework of how CO2 and O3 affect belowground C cycling. We briefly summarize the changes in aboveground productivity and then follow the path of C belowground, from fine roots and mycorrhizae to free-living soil microorganisms and soil respiration. We conclude by returning to the conceptual framework and putting the results into a broader scientific context.
10.2.1 Net Primary Productivity and Community Composition Over the first 3 years of the experiment, these forest communities responded to CO2 and O3 as predicted: elevated CO2 increased net primary productivity
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(NPP), while elevated O3 reduced NPP (King et al., 2005). These changes were consistent with leaf-level measurements of light-saturated photosynthesis, which were higher under elevated CO2 and lower under elevated O3 (Darbah et al., 2010). However, the effects of elevated O3 were smaller and less consistent than the effects of elevated CO2 (Kets et al., 2010). Interestingly, the negative effect of O3 on NPP began to diminish after 3–5 years (depending on the community; King et al., 2005) and during the final 3 years of the experiment, elevated O3 no longer reduced NPP (Zak et al., 2011). The recovery in NPP may be a consequence of compensatory growth of O3tolerant plants in response to the suppression of O3-sensitive species and genotypes (Zak et al., 2011). For instance, exposure to elevated O3 and ambient CO2 increased the relative importance of birch and maple over aspen (Kubiske et al., 2007). However, the effect of elevated O3 on leaf litter production (g m2) remained relatively constant over the last 7 years of the experiment, averaging 13% (Talhelm et al., 2012). In contrast, elevated CO2 had a sustained positive effect on NPP (þ25% in 2008; Zak et al., 2011) and leaf litter production (þ32% in 2008; Talhelm et al., 2012).
10.2.2 Fine Roots Dynamics In these developing forests, fine root (<1 mm in diameter) biomass increased quickly during the first half of the experiment before stabilizing over the last 4 years of the experiment (Figure 10.2). Within this overall pattern, the response of fine root biomass to the treatments was highly dynamic. In the
FIGURE 10.2 Fine root (<1 mm in diameter) biomass during the Aspen FACE experiment. Data from estimates in King et al. (2005), Pregitzer et al. (2006, 2008), and Zak et al. (2011) for the top 25 cm of mineral soil. Circles represent ambient O3, triangles represent elevated O3, empty symbols represent ambient CO2, and filled symbols represent elevated CO2. Error bars are 1 standard error.
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initial sampling (1999), elevated CO2 nearly doubled fine root biomass in the top 10 cm of mineral soil, but elevated O3 did not have a significant effect (King et al., 2001). In subsequent samplings in 2000 and 2002 (to 25 cm in depth; Figure 10.2), elevated CO2 created increases in fine root biomass that were smaller but consistent across all community types (averaging þ37%), while the response to elevated O3 was community-specific (King et al., 2005; Pregitzer et al., 2008; Figure 10.3). Elevated O3 decreased fine root biomass within the aspen-birch and aspen-maple communities in these two samplings, but O3 unexpectedly stimulated fine root biomass by more than 30% within the aspen-only community (King et al., 2005; Pregitzer et al., 2008). The response of fine root biomass to elevated CO2 was still positive in 2005 and 2008 (Zak et al., 2011), but this stimulation shrank further to 27% in 2005 and 11% in 2008. The effect of elevated O3 was positive across all communities in 2008 (þ4%), but the stimulation of fine root biomass in the aspen-only community was smaller (þ9%) than in previous estimates. Less is known about the response of roots deeper in the soil. However, Rhea and King (2012) sampled fine roots in the aspen-only and aspen-birch communities to 1 m depth in 2005 and found that neither treatment significantly affected fine root biomass below the top 30 cm of soil. It is interesting to contrast leaves and fine roots, the two most ephemeral major elements of plant biomass. Like roots, leaves initially displayed both rapid growth and large treatment effects, followed by a period of more subtle growth and smaller treatment effects (King et al., 2005; Talhelm et al., 2012). Leaf and fine root biomass both responded positively to elevated CO2 in all three communities and negatively to O3 in the aspen-birch and aspen-maple communities (Figure 10.2), but the positive response of fine roots to elevated O3 in the aspen community was not observed for leaves (Figure 10.2). Although this divergent response was unique among the three communities, the negative response to elevated O3 was consistently stronger for leaf biomass than for fine root biomass in all three communities (Figure 10.3). Similarly, the ratio of leaf area to fine root area had a positive, though less consistent, response to O3 when measured in 2005 (Rhea and King, 2012). The production and mortality of fine roots in the aspen community was estimated using minirhizotron observations from 2002 to 2005 (Pregitzer et al., 2008). Production rates (mm mm1) were not affected by CO2 or O3, but elevated O3 significantly increased root mortality rates in 1 year (2003). Consequently, changes in production and mortality biomass (g m2 year1) were mostly a function of changes in standing biomass (Pregitzer et al., 2008). These responses contrast with observations of leaf life span in the aspen-only community, in which leaf abscission was delayed by elevated CO2 and unaffected by O3 (Riikonen et al., 2008; Taylor et al., 2008).
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FIGURE 10.3 Relative treatment effect size (elevated/ambient) for leaf biomass and fine root (<1 mm in diameter) biomass for each community. Leaf biomass effects from King et al. (2005) and Talhelm et al. (2012); fine root data sources as in Figure 10.2. Triangles represent fine roots, circles represent leaves, empty symbols represent O3 effects, and filled symbols represent CO2 effects. Horizontal lines represent no treatment response.
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10.2.3 Mycorrhizal Fungi There are no estimates of the mycorrhizal biomass or hyphal turnover rates. However, insight into the influence of CO2 and O3 on mycorrhizae is available from observations of ectomycorrhizal sporocarps (2003–2006) within the aspen-only and aspen-birch communities (Andrew and Lilleskov, 2009) and analyses of fungal ribosomal DNA within the mineral soil in 2002 (Lesaulnier et al., 2008) and 2007 (Edwards and Zak, 2011). Elevated CO2 increased overall sporocarp production by an average of 96% in the aspenonly community and 70% in the aspen-birch community, principally by increasing the abundance of late-successional ectomycorrhizal taxa. Elevated CO2 also increased the species richness of these sporocarps. Notably, elevated CO2 increased the abundance of the early-successional ectomycorrhizal taxa Inocybe in the 2002 molecular analyses (Lesaulnier et al., 2008), but this change was not evident in 2007 (Edwards and Zak, 2011). Elevated O3 did not affect sporocarp species richness and only affected sporocarp production and community composition under ambient CO2 (Andrew and Lilleskov, 2009). Here, elevated O3 decreased sporocarp production by an average of 12% in the aspen-birch community and, in contrast to the response of fine root biomass, decreased sporocarp production by 48% in the aspen-only community. Elevated O3 affected sporocarp community composition under ambient CO2 by increasing the abundance of earlysuccessional taxa such as Inocybe and Paxillus. Assessments of fungal community composition in 2007 based on ribosomal DNA abundance provided more evidence that elevated O3 had increased the abundance of Inocybe, but in general, O3 effects on ectomycorrhizal community structure in these molecular analyses were small and independent of CO2 (Edwards and Zak, 2011). In addition to Inocybe, ectomycorrhizal communities within the soil under elevated O3 had relatively more species of Cortinarius and fewer Laccaria and Tomentella species (Edwards and Zak, 2011). Although Andrew and Lilleskov (2009) found more early-successional species under elevated O3 and more late-successional species under elevated CO2, there were clear differences in sporocarp community composition between the ontogenetically similar ambient and þCO2 þ O3 treatments.
10.2.4 Soil Respiration For the first 7 years of the experiment, soil respiration was measured only in the two most productive communities, aspen-birch and aspen-only. Elevated CO2 stimulated soil respiration in these two communities throughout the first seven growing seasons of the experiment (1998–2005), but the size of this stimulation varied by community, year, and season (King et al., 2001, 2004; Pregitzer et al., 2006). For instance, the annual CO2 effect during the first 4 years ranged from þ13% to þ41% within the aspen-only community and
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from þ43% to þ77% within the aspen-birch community (King et al., 2004). Elevated CO2 continued to have a smaller effect on soil respiration in the aspen-only community than the aspen-birch community throughout the first decade of the experiment (Pregitzer et al., 2006, 2008) and effect of elevated CO2 was not significant within the aspen-only community in 2002. Here, soil respiration rates were similar in the aspen-only and aspen-birch communities under elevated CO2, but soil respiration tended to be higher under ambient CO2 in the aspen-only community (Pregitzer et al., 2006, 2008). Later in the experiment (2005–2007), soil respiration was also quantified within the aspen-maple community. Patterns in growing season soil respiration in this community were similar to those in the aspen-birch community in terms of the size of the overall soil CO2 flux and the relative effects of elevated CO2. During these 3 years, the annual stimulation across all three communities of growing seasonal soil respiration by elevated CO2 ranged from 25% to 31% (Pregitzer et al., 2008). In general, the magnitude of the CO2 stimulation varied seasonally; it was strongest in the middle of the growing season when soil respiration was greatest and weaker during the spring and fall when soil temperatures were low (King et al., 2001; Pregitzer et al., 2006). Elevated O3 had much less consistent effects on soil respiration. There was no significant O3 effect on soil respiration in 1998, but in 1999 elevated O3 caused a 20% decrease in soil respiration late in the growing season (August to November; King et al., 2001). Elevated O3 significantly decreased soil respiration within the aspen-only community in 2002, but in the two subsequent years there was a significant CO2 O3 interaction (P < 0.01) in both communities and the effect of elevated O3 became positive under elevated CO2 (Pregitzer et al., 2006). Rates of soil respiration were consistently 5–10% greater in the þCO2 þ O3 treatment than in the þCO2 treatment from 2005 to 2007, but the CO2 O3 interaction was not significant because elevated O3 often only slightly decreased soil respiration under ambient CO2. This slight positive effect of O3 was most consistent in the aspen-only community, matching the response of fine root biomass (Pregitzer et al., 2008). In fact, soil respiration rates in 2005 were significantly correlated with fine root biomass across treatments and communities (r ¼ 0.72; P ¼ 0.008; Pregitzer et al., 2008). In addition to increasing the flux of CO2 from the soil, elevated CO2 also increased the partial pressure of CO2 within the soil (Karberg et al., 2005; King et al., 2001). The increase appears to have resulted primarily from greater rhizosphere C fluxes rather than changes in soil moisture or temperature that would reduce the diffusion of CO2 (Karberg et al., 2005; King et al., 2001). Additional CO2 within the soil is important because it can dissolve into the soil solution and increase the rate of two key processes: soil weathering and long-term C sequestration within oceanic carbonates (Kump et al., 2000). In measurements throughout the 2002 growing season, Karberg et al.
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(2005) observed that increases in the partial pressure of soil CO2 resulted in a 22% increase in dissolved inorganic carbon (DIC) and a 14% increase in the concentration of carbonic acid in soil solution. Soil solution alkalinity also increased by 210% (Karberg et al., 2005), likely due to an enhancement of primary mineral weathering by carbonic acid. Although elevated O3 did not affect either the soil partial pressure of CO2 or the concentration of DIC, we note that DIC concentrations in 2002 followed the same rank order as in soil respiration measurements made in subsequent years (2003–2004; Pregitzer et al., 2006): þCO2 þ O3 > þ CO2 > ambient > þ O3. In contrast to the changes in DIC, measurements of dissolved organic carbon during the first 2 years of the experiments found no clear effect of elevated CO2 or O3 (King et al., 2001).
10.2.5 Soil Organic Matter In addition to changing the amount of fine root and leaf litter, CO2 and O3 also influenced the chemical composition of plant litter and low molecularweight organic compounds in the surface mineral soil. In the mineral soil, exposure to elevated CO2 and ambient O3 created a marginally significant (P ¼ 0.08) increase in sugar concentrations (approximately þ10%) and had a larger effect on soluble phenolics (P ¼ 0.07, approximately þ40%), but there were no other treatment effects on sugar or soluble phenolic concentrations (Johnson and Pregitzer, 2007). These differences contrast with changes in plant litter chemistry. Concentrations of non-structural carbohydrates (sugars and starch) and structural carbohydrates (cellulose and hemi-cellulose) in fine roots collected in 2002 were not affected by the treatments (Chapman et al., 2005). Concentrations of sugars and soluble phenolics in leaf litter increased in response to CO2 (Liu et al., 2005; Parsons et al., 2008), but the greatest increases tended to be in the interaction (þCO2 þ O3) treatment rather than with exposure to elevated CO2 alone (þCO2). Soil amino acid concentrations responded individualistically to CO2 and to O3. Elevated CO2 increased aspartic acid concentrations, while elevated O3 decreased arginine and methionine concentrations. Elevated O3 decreased valine concentrations under ambient CO2, but had no effect under elevated CO2. The 11 other amino acids assayed were unaffected by the treatments (Johnson and Pregitzer, 2007). Averaged across all communities and treatments, soil C content within the top 20 cm of mineral soil increased from 3350 230 g m2 when first measured in 2001 to 4010 160 g m2 in 2008 (Talhelm et al., 2009). Although elevated CO2 consistently increased NPP (King et al., 2005; Zak et al., 2011), fine root production, and leaf litter inputs (Talhelm et al., 2012), there was no increase in soil C content (Figure 10.3). Soil C content was not significantly affected by elevated CO2 in the aspen-birch and aspen-maple communities. In the aspen-only community, soil C accrued more slowly under elevated CO2 than under ambient CO2 during the last several
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years of the experiment, creating a large and significant negative effect (1740 g m2) of elevated CO2 during the final year of the experiment (P ¼ 0.003; Figure 10.3). Likewise, although elevated O3 reduced NPP early in the experiment, it did not affect soil C content. The mass of the soil organic horizon (forest floor) was somewhat greater under elevated CO2 and smaller under elevated O3 in both 2004 and 2008, but these differences were not significant (Zak et al., 2011). The lack of soil C accumulation under elevated CO2 contrasts with studies of leaf litter and fine root decomposition (Chapman et al., 2005; Liu et al., 2009; Parsons et al., 2008), wherein elevated CO2 either slowed decomposition (Chapman et al., 2005; Parsons et al., 2008) or had no significant effect (Liu et al., 2009). For elevated O3, litter decomposition studies had mixed results: increased decomposition (Parsons et al., 2008), decreased decomposition (Liu et al., 2009; Parsons et al., 2008), or no effect (Chapman et al., 2005). Insight into the changes in C cycling within the mineral soil can be gained by isolating and quantifying different fractions of soil organic matter. Hofmockel et al. (2011) partitioned soil organic matter sampled from the top 20 cm of mineral soil in 2003, 2004, and 2007 into several physical fractions: coarse particulate (>250 mm), fine particulate (250–53 mm), and mineral-associated (<53 mm). The majority of soil C was found in the mineral-associated fraction, nearly 3000 g m2. The pools of coarse particulate and fine particulate organic matter were considerably smaller, with each accounting for approximately 500 g m2 of C. Although elevated CO2 did not have significant effects on any of these fractions overall, it did affect how the pools of coarse particulate and mineral-associated soil C changed through time. Coarse particulate C pools increased faster under elevated CO2 and were 19% greater under elevated CO2 in 2007. However, this increase was matched by a trend towards decreased mineral-associated C as the effect of elevated CO2 changed from þ7% in 2003 to 1% in 2007. Together, these trends suggest that while there is no significant overall effect on soil C content, the processing of soil C has been altered by elevated CO2 (Hofmockel et al., 2011). The effects of elevated CO2 on coarse particulate soil C also varied significantly by community, with a 6% decrease within the aspen-only community and a 23% increase in the aspen-birch community. There was a similar pattern in the pool of mineral-associated soil C, but these differences were not significant. Elevated O3 did not significantly affect the amount of C in these three soil fractions, but did increase the C/N ratio and decrease the pool of soil N within the coarse particulate and fine particulate fractions. The fossil-fuel derived CO2 used to create the elevated CO2 treatment is heavily depleted in 13C. This depletion is reflected in the d13C of plant tissue and creates an opportunity to distinguish ‘new’ C added to the soil since the beginning of the fumigation and ‘old’ pre-existing C. In the initial survey of these pools within the aspen-only and aspen-birch communities in 2001, the
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amount of total new soil C formed under elevated CO2 was 51% lower under elevated O3 (þCO2 þ O3 vs. þCO2; Loya et al., 2003). Using a chemical fractionation technique, Loya et al. (2003) also found that approximately onethird of this new C was recalcitrant acid-insoluble organic matter and that elevated O3 reduced the formation of new acid-insoluble C by 48%. When these soils were incubated, more C was respired in the interaction (þCO2 þ O3) treatment than under elevated CO2 alone, suggesting that the changes in soil C pools were a product of increased C losses rather than decreased C inputs (Loya et al., 2003). However, the negative effect of elevated O3 on soil C formation under elevated CO2 was considerably weaker within a wider analysis that included all three communities and measurements from 2001 to 2008 (Talhelm et al., 2009). Elevated O3 did not significantly decrease the amount of new soil C formed under elevated CO2 in the aspen-maple community in 2001 and the negative effects of O3 within the aspen-birch and aspen-maple communities weakened in subsequent measurements and were no longer significant. The input of new soil C under elevated CO2 was evident within each of the three physical soil C fractions, but inputs of new C to the fine particulate and mineral-associated soil C were offset by loss of old soil C (Hofmockel et al., 2011). Elevated O3 reduced the fraction of new C somewhat within each of the three fractions, but not significantly. However, elevated O3 slightly, but significantly, increased the amount of old C within the coarse particulate and fine particulate soil C pools (Hofmockel et al., 2011). Although this isotopic technique cannot be used to determine why elevated CO2 reduced total soil C content within the aspen-only community, it is notable that the largest loss of pre-existing ‘old’ soil C under elevated CO2 occurred within this community (Talhelm et al., 2009).
10.2.6 Soil Microorganisms and Extracellular Enzymes The lack of soil C accumulation under elevated CO2 was surprising, but consistent with observations of soil extracellular enzymes. Cellobiohydrolase and N-acetylglucosaminidase are extracellular enzymes important in the microbial breakdown of cellulose and chitin, respectively, which are two of the most abundant biomolecules within soil. Early in the experiment (1999), when the relative effects of elevated CO2 on fine roots and leaf litter were large, elevated CO2 stimulated the activity of cellobiohydrolase and N-acetylglucosaminidase by approximately 50% (Larson et al., 2002). A similar stimulation of these enzymes by elevated CO2 was also observed in 2001 and 2002 (Chung et al., 2006). These observations were consistent with laboratory incubations in which elevated CO2 increased the respiration of added 13C-labelled substrates for these enzymes (Phillips et al., 2002). Interestingly, the effect of elevated CO2 on cellobiohydrolase was considerably stronger under ambient O3 than under elevated O3 in both the field and the laboratory incubation, but this was not apparent for N-acetylglucosaminidase
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(Edwards and Zak, 2011; Phillips et al., 2002). The activities of other soil extracellular enzymes used to breakdown soil organic matter, including phosphatase, peroxidase, 1,4-a-glucosidase, 1,4-b-glucosidase, 1,4-b-xylosidase, and1,4-b-N-acetylglucose-aminidase also tended to be higher under elevated CO2 early in the experiment, but these differences were sometimes not significant (Chung et al., 2006; Larson et al., 2002). As the relative effects of elevated CO2 on fine roots and leaf litter declined, the stimulation of cellobiohydrolase and N-acetylglucosaminidase by elevated CO2 also declined and was no longer significant by 2008 (Edwards and Zak, 2011). The shrinking effect of elevated CO2 on N-acetylglucosaminidase resulted from an increase in enzyme activity under ambient CO2 rather than a decrease under elevated CO2, but the changes through time in cellobiohydrolase were less consistent (Edwards and Zak, 2011). Elevated O3 had a smaller impact on extracellular enzymes. Elevated O3 reduced cellobiohydrolase activity within the Oe horizon by 17% in 2008 (Edwards and Zak, 2011) and reduced the activity of 1,4-b-glucosidase by 25% in the surface mineral soil in 2001–2002 (Chung et al., 2006). Matching the inconsistent positive response of some enzymes to CO2, a similar group of enzymes tended to decrease under elevated O3. Elevated CO2 and elevated O3 each increased microbial (Larson et al., 2002) and fungal biomass (Chung et al., 2006), but not significantly. Although plant community type and soil horizon appear to be the dominant controls, elevated CO2 and O3 did influence microbial community composition within the mineral and organic soil (Chung et al., 2006; Edwards and Zak, 2011). For instance, elevated CO2 appears to increase the abundance of Sistotrema fungi, while elevated O3 shifted the composition of agaricomycete fungi (Edwards and Zak, 2011). Elevated CO2 also increased bacterial taxonomic richness (Dunbar et al., 2012; Lesaulnier et al., 2008), but decreased archaeal diversity and reduced the abundance of archaeal DNA by 50% (Lesaulnier et al., 2008). There is evidence that elevated CO2 altered the abundance of a number of individual bacterial and archaeal taxa (Dunbar et al., 2012; Lesaulnier et al., 2008). For instance, of 1313 individual bacterial taxa identified at Aspen FACE, 75 taxa significantly increased or decreased in abundance in response to elevated CO2 (Dunbar et al., 2012). However, the functional roles of individual taxa are poorly defined (Dunbar et al., 2012).
10.3 CONCLUSIONS AND IMPLICATIONS Belowground C cycling responded dynamically to the CO2 and O3 treatments within the developing forests at Aspen FACE. Often, these processes were consistent with changes in aboveground productivity (Figure 10.3). For instance, elevated CO2 increased photosynthesis and aboveground productivity, resulting in concomitant increases in fine root biomass, soil respiration, the partial pressure of soil CO2 and DIC. However, some responses diverged
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from expectations in several important ways, particularly for O3 (Figure 10.1). First, elevated O3 reduced leaf biomass more than fine root biomass in all communities, contrasting with reviews that have found decreased root:shoot ratios under elevated O3 (Anderson, 2003; Grantz et al., 2006; Wittig et al., 2009). Further, elevated O3 stimulated fine root biomass in the aspen-only community and stimulated soil respiration in all communities during the latter part of the experiment. A positive response of both fine root biomass and soil respiration was found in another free-air forest O3 experiment (Nikolova et al., 2010; Matyssek et al., 2013, this vol.) and, in general, the response of fine root biomass to O3 is highly variable in developing forests (King et al., this vol.). Likewise, mycorrhizal sporocarp production responded differently to CO2 or O3 than did fine roots, sometimes in the opposite direction. Finally, changes in productivity were not reflected in soil C content (Figure 10.4), as extracellular enzyme activity changed in parallel with litter inputs.
FIGURE 10.4 Relative treatment effect sizes for soil carbon content for each community. Empty symbols represent O3 effects and filled symbols represent CO2 effects. Significant (P < 0.05) CO2 effect denoted with asterisk.
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Overall, the differences between the predicted and observed effects on belowground processes illustrate the gaps in our knowledge of these processes and the value in long-term field experiments. Unfortunately, similar experiments (in scale and duration) are too few in number to know if the responses of belowground processes at Aspen FACE are universal or exceptional. Because of the large role of belowground processes in terrestrial C cycling (Bond-Lamberty and Thomson, 2010; Litton et al., 2007; Pregitzer and Euskirchen, 2004), accurate predictions of future terrestrial productivity and C sequestration will require further long-term experiments to achieve a more sophisticated understanding of how these processes respond to air pollution, climate change, and other global change factors. Without accurate predictions of future forest C cycling, the ability to design effective climate policy will be limited.
ACKNOWLEDGEMENTS The Aspen FACE experiment was supported by funding from the U.S. Department of Energy and by the USDA Forest Service. We sincerely thank Donald Zak, Dave Karnosky, Andrew Burton, Jud Isebrands, George Hendrey, Richard Dickson, John Nagy, and Keith Lewin for their invaluable contributions.
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