Effects of climate change on leaf litter decomposition across post-fire plant regenerative groups

Environmental and Experimental Botany 77 (2012) 274–282

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Effects of climate change on leaf litter decomposition across post-fire plant regenerative groups b ˜ S. Saura-Mas a,∗ , M. Estiarte b , J. Penuelas , F. Lloret a a CREAF (Center for Ecological Research and Forestry Applications) and Unit of Ecology, Department of Animal and Plant Biology and Ecology, Autonomous University of Barcelona, E-08193 Bellaterra, Barcelona, Spain b Global Ecology Unit CREAF-CEAB-CSIC, CREAF (Center for Ecological Research and Forestry Applications), Edifici C, Universitat Autònoma de Barcelona, 08193, Bellaterra, Spain

a r t i c l e

i n f o

Article history: Received 20 January 2011 Received in revised form 27 October 2011 Accepted 19 November 2011 Keywords: Decomposition rate Resprouting Warming Drought Leaf litter Mediterranean type-ecosystems

a b s t r a c t Decomposition is a determining factor for the functioning of ecosystems because litter dynamics (litter fall and litter decomposition) constitute a key process in the regulation of the recycling of carbon and nutrients. We studied the litter decomposition properties of a set of 19 Mediterranean-basin woody species with different post-fire regenerative strategies (resprouters and non-resprouters), under experimental climate manipulation (simulating warming and drought) over a 2-year period. We show that climate change modifies litter decomposition of these Mediterranean woody species as litter contributions to the soil (g/year) were lower under drought experimental conditions. Species with different post-fire regeneration performance showed different leaf decomposition patterns, though these patterns were influenced by the taxonomical affiliation of the species. As expected, the mass loss of the non-resprouter litter, after 2 years, was higher than in resprouters. Non-resprouters showed higher nutrient concentration per mass of leaf litter after 2 years of experiment than resprouters, possibly because they have lost more C-rich biomass, allowing high nutrients concentration in the remaining litter. That would apply particularly to P as litter N:P ratio was lower in non-resprouters than in resprouters. This study suggests that, in Mediterranean ecosystems, nutrients’ return from leaf litter to the soil will be slower under the projected future drier conditions. Furthermore, changes in fire regimes that lead to modifications in the abundance of post-fire regenerative groups are likely to affect ecosystem’s functional properties. Thus, if new fire regimes enhance non-resprouters’ abundance, we can expect a greater return of organic matter contributions to the soil and a lower litter N:P. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Shrublands occupy extensive areas of the Mediterranean region (Moreno and Oechel, 1995; Riera et al., 2007), where the availability of water is a key factor determining vegetation composition and patterns of plant distribution. A trend to progressive aridification has been seen over the last few decades in many of these areas, as a result of increased evapotranspiration (caused by approximately 1 ◦ C average warming), without any parallel increase in rainfall ˜ ˜ ˜ et al., 1998; Penuelas et al., 2002; Penuelas and Boada, 2003). (Pinol Furthermore, climatic projections anticipate further warming and aridification over the coming decades in the Mediterranean region ˜ (Penuelas et al., 2005; IPCC, 2007). Apart from aridity, drought and warming, it also increases the risk of wildfires. In fact, the frequency of wildfires in some areas of the Mediterranean basin has increased over the last few decades

∗ Corresponding author. E-mail address: [email protected] (S. Saura-Mas). 0098-8472/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.envexpbot.2011.11.014

as a result of climate change and changes in land use and human ˜ et al., 1998; Pausas, 2004). Evosocioeconomic activities (Pinol lution and dynamics of most Mediterranean-type ecosystems are linked to wildfires (Hanes, 1971; Whelan, 1995; Lloret et al., 2002; Keeley et al., 2011; Bradshaw et al., 2011) and most Mediterranean woody species present regeneration mechanisms after this type of disturbances, determining several post-fire regeneration types. Obligate seeders germinate abundantly after fire from soil or canopy banks, counterbalancing fire-induced mortality. Obligate resprouters diminish fire-induced mortality thanks to vegetative organs with some kind of protection against high temperatures that are able to resprout after wildfires. Seeder–resprouters (or facultative seeders) exhibit both mechanisms, and finally a few species – at least in Mediterranean ecosystems – are not able regenerate well after wildfire. The combination of these responses results on four groups of species (sensu Pausas and Verdú, 2005): seeders (S+ R−), resprouters (S− R+), seeders–resprouters (S+ R+) and those that do not regenerate after wildfire (S− R−). The proportion of these types of species in plant community depends, among other factors, on fire regime and climatic gradients. In the Mediterranean

S. Saura-Mas et al. / Environmental and Experimental Botany 77 (2012) 274–282

basin there is a trend toward an increased proportion of short-lived seeders under moderately high fire frequencies (Lloret et al., 2005; Vila-Cabrera et al., 2008). Also in this region, non-resprouters (S+ R−) have a shorter life cycle and a higher recruitment rate after disturbances such as fire or drought (Verdaguer and Ojeda, 2002; Pausas and Verdú, 2005). In contrast, the group of resprouters (S− R+, S+ R+) is mostly constituted by long-lived species, typically those found in late successional shrublands, in which a high percentage of individuals survive and resprout after fire but do not present any significant short-term recruitment of new individuals. So, long-lived resprouters are expected to show a higher resource allocation to underground organs in order to sustain plant re-growth (Pate et al., 1990; Knox and Clarke, 2005; Schwilk and Ackerley, 2005), while short-lived non-resprouters would have a shorter leaf lifespan and higher photosynthetic rates (Bell, 2001; Ackerly, 2004), similar to early successional species of mesic forests (Bazzaz, 1979). These traits are expected to be reflected in the leaf litter decomposition of these groups of species. These differences could be heightened by differences in leaf properties, such as lower leaf dry-matter content and higher seasonal re-hydration capacity in seeders versus resprouters (Saura-Mas and Lloret, 2007). Decomposition is a key to better understanding of the effects of climate change, particularly those involving growth as well as carbon and nutrient cycles. It is thus a determining factor for ecosystem production and there is a significant positive lineal relationship between the decomposition rate and relative growth rate for different types of communities, including shrublands (Cebrián and Duarte, 1994, 1995; McTiernan et al., 1997; Maisto et al., 2011). Moreover, the balance between net primary production and decomposition strongly influences carbon and nutrient cycling at ecosystem scale (Chapin et al., 2002). Litter dynamics (litter fall and litter decomposition) constitute a key process in the functioning of ecosystems, as they exert a decisive effect on the recycling of carbon and nutrients. Litter fall represents an output of nutrients from the aerial parts of the plants, and also an input of nutrients to the soil. Subsequent decomposition is the route by which some of the carbon fixed by plants and nutrients is partially incorporated into the decomposing biomass, as inorganic nutrients in the soil or returned to the atmosphere as CO2 . So, this process releases carbon into the atmosphere, as well as nutrients in forms that can be used for plant and microbial production (Chapin et al., 2002; Gartner and Cardon, 2004). This conversion of dead organic matter (leaf litter) occurs by means of a leaching action (removing soluble materials from decomposing organic matter), fragmentation (by soil-dwelling animals that break large pieces of organic matter into smaller ones and mix the decomposing organic matter into the soil) and chemical alteration (primarily through the activity of bacteria and fungi). The decomposition rate is therefore regulated by a set of factors that affect soil biota activity: physical environment (mainly the climate), litter composition and substrate nutrients (Kang et al., 2010; Prescott, 2010). In this study our main objective was to explore the differences in litter decomposition between species of Mediterranean basin shrublands with different post-fire regeneration under an experimental climatic manipulation of temperature and rainfall. Beyond this, our objective was to establish a link between ecosystem functioning and community composition defined by species attributes related to post-disturbance regeneration, according to their sensitivity to current trends of climate change (i.e. decrease in water availability, increase of temperatures and increase in fire occurrence). More specifically, we tested two main hypotheses. First of all, we hypothesized that since resprouter species have higher leaf dry-matter content (Saura-Mas and Lloret, 2007) than non-resprouters, then, leaf litter of these species would present a lower leaf litter decomposition likely promoting a slower flux of carbon to the soil. The second hypothesis was that higher

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temperatures and drought could alter soil biota species composition and metabolic activity, affecting litter decomposition rate. However these effects may be complex because in the meantime that temperature and soil humidity are related, higher temperatures tend to increase metabolic activity, while low soil humidity should deplete microbial functioning. In these assessments, we also considered the taxonomical affiliation of species, as differences between the functional properties of post-fire regenerative groups could be linked to their evolutionary history (Verdú, 2000; Pausas and Verdú, 2005; Saura-Mas and Lloret, 2009b; Verdú et al., 2007). 2. Materials and methods 2.1. Species sampling The study was carried out on a subset of 19 woody plant species growing on limestone in coastal shrublands of Catalonia (northeast of the Iberian peninsula). They were representative of a community that is widely distributed in the region and belonged to as many different families, life-form types and post-fire regenerative groups as possible (Table 1). Depending on their post-fire regenerative strategy (Cucó, 1987; Papió, 1988; Lloret and Vilà, 1997; Verdú, 2000; Alberdi and Cavero, 2003; Lloret et al., 2003; Paula et al., 2009), and after direct field observations in a nearby area that was burnt in September 2004, species were classified into two post-fire regenerative groups: resprouters (R: S− R+, S+ R+) and non-resprouters (S+ R−). Freshly, senescent leaves were collected from the plants of the different species from March 2003 to March 2004. The precise time of collection depended on their phenology (Floret et al., 1989). Leaf litter was collected in the Montgrí Massís (except for Arbutus unedo and Globularia alypum, which were collected in Garraf), a protected coastal area located in the NE of Catalonia (northeast Iberian Peninsula, 42.16◦ N, 3.24◦ W). Vegetation is mainly constituted by open Pinus halepensis forests and shrublands, dominated by Quercus coccifera, Cistus albidus, Cistus monspeliensis, and Rosmarinus officinalis (Polo and Masip, 1987). The annual precipitation is 655 mm, with cool winters and warm summers (mean annual temperature: 14.8 ◦ C) (Ninyerola et al., 2000, 2003). Sampling was conducted in mature shrubland (1–2 m high) that had been untouched by wildfire for over 10 years. The Montgrí community has been sampled for the characterization of different leaf attributes (Saura-Mas and Lloret, 2007) of the main species, while Garraf was the nearest site with the same type of ecosystem (coastal shrubland on limestone) and an estab˜ lished experimental setting of climate manipulation (Penuelas et al., 2004). The experiment was therefore located in Garraf, a protected coastal area located in the NE of Catalonia (41.19◦ N, 1.49◦ W). The vegetation is similar to the Montgrí shrubland and is mainly dominated by Q. coccifera, Globularia alypum, Erica multiflora and R. officinalis. The annual precipitation is 455 mm, with cool winters and warm summers (mean annual temperature: 15.1 ◦ C). The study area soil type is described as a petrocalcic calcixerept soil (pH 8.2), with low quantities of organic matter in the upper soil horizons (3.5%). 2.2. Experimental design 2.2.1. Climate treatments Nine experimental plots (5 m × 4 m) were established in relatively homogeneous areas within the Garraf study area. Three treatments were allocated: control (C), warming (W), and prolonged drought during the growing season (D). A light scaffolding structure was built around each plot, comprising galvanized steel tubes covered by thin plastic sleeves to prevent contaminants from

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leaching into the plot. In the warming plots, this frame supports a retractable, reflective curtain made of strips of infrared-reflective material bound into a high-density polyethylene mesh. A small motor activated by a light sensor extends this curtain over the vegetation at night, thus preventing heat loss. In this warming treatment, a tipping-bucket rain sensor retracts the curtain at night to enable rain to enter the plot. To prevent damage to the cover, a wind sensor activates the retraction of the curtain at night if wind speeds exceed 10 m/s. On the drought plots, a retractable curtain made of transparent polyethylene plastic prevents rainfall during the main growing seasons, between March and June in spring, and between September and December in autumn. Rain sensors activate the motor to extend this cover over the plots when rain is detected, and to retract the cover when the rain stops. When the curtain is extended, wind sensors again retract the curtain to prevent damage during periods of high wind. This climate change experiment has been in operation since 1999 (Beier et al., ˜ 2004; Penuelas et al., 2007). The effect of the warming treatments on mean air temperature (measured at 20 cm above the soil) was 0.59 ◦ C in 2004, 0.68 ◦ C in 2005 and 0.68 ◦ C in 2006 (treatments in 2006 are considered from January to September, when the experiment was completed). The effect of the warming treatments on mean soil temperature (measured at 5 cm depth from the surface of the soil) was 0.66 ◦ C in 2004, 0.78 ◦ C in 2005 and 0.60 ◦ C in 2006. The growing-season drought treatment reduced precipitation by 48.87% (2004, 155 days of treatment), 47.21% (2005, 154 days of treatment), and 4.21% (2006, 33 days of treatment) (Fig. A.1). The reached temperatures in the warming treatment are in consonance with the increase of the mean annual temperature of the region, which has been estimated to be approximately between 0.18 ◦ C and 1 ◦ C in a decade ˜ et al., 1998; Penuelas ˜ ˜ (Pinol et al., 2002; Penuelas and Boada, 2003; Martín-Vide et al., 2010). All samplings and measurements were made in an inner plot measuring 4.5 m × 3.5 m plot, avoiding an outer 0.5 m buffer to obviate any edge effects.

decomposition bags measured 10 cm × 9 cm and were made of 0.5 mm nylon mesh. The mesh size was small enough to prevent any major losses of the smallest leaves, and yet large enough to permit microbial and fungi activity, as well as the free access of small soil animals (Killham, 1994; Lavelle, 1996). Between 1 and 2 g of air-dried leaf litter of the 19 woody species (Table 1) was added to each handmade nylon mesh bag and the resulting mass was recorded. The bags were sealed with stainless steel staples and labeled with aluminum tags. Two decomposition experiments were performed, one within plots with climatic treatment (hereafter climatic experiment) and another outside the treated plots (hereafter non-treated experiment). The latter experiment was performed to obtain additional information on the decomposition of each species over time. The experiments were set up on 6th September 2004 and finished after 2 years, on 6th September 2006. A total of 608 litter bags were placed in the study area. For the climatic experiment, we placed, on randomly selected sites, two bags for each species (except for Daphne gnidium and Fumana thymifolia, because of the lack of sufficient leaf litter) in each of the nine experimental plots. One bag per species and plot were collected after 0.17 and 2 years of incubation (3 replicates per species, treatment and time). For the non-treated experiment, we randomly placed 16 litter bags per species outside the treatment plots and four replicates per species were collected after 0.17, 0.5, 1.17, 2 years of incubation. F. thymifolia had only three replicates, because it was not possible to obtain enough material, due to the small plant and leaf size.

2.2.2. Field litter decomposition Decomposition experiments were carried out using the technique of the mesh bag placed over the soil surface (Gallardo and Merino, 1993; Murphy et al., 1998; Kazakou et al., 2006). The

Each individual sample was ground and analyzed for nitrogen (N), phosphorous (P), calcium (Ca), potassium (K), sodium (Na) and magnesium (Mg). Nutrient concentrations are expressed as mg of nutrient per g of leaf litter biomass, and nutrient contents

2.2.3. Laboratory analysis The samples were oven-dried at 60 ◦ C for 24 h immediately after their return from the field. After the removal of litter from the bags, the remaining litter was weighed to the nearest 0.0001 g and % of remaining mass was calculated as: Remaining mass × 100. Initial mass

Table 1 Study species, family, post-fire regenerative strategies and life-form (according to Raunkiaer classification). Chamaephyte = C, nano-phanaerophyte = NP, macrophanaerophyte = MP, panaerophyte-vine = PV. Study species

Arbutus unedo L. Cistus albidus L. Cistus monspeliensis L. Cistus salviifolius L. Daphne gnidium L.a Dorycnium hirsutum (L.) ser. In DC. Dorycnium pentaphyllum Scop. Fumana thymifolia L.1 Globularia alypum L. Juniperus oxycedrus L. Lavandula latifolia Med. Phillyrea angustifolia L. Pistacia lentiscus L. Quercus coccifera L. Quercus ilex L. Rhamnus alaternus L. Rosmarinus officinalis L. Smilax aspera L. Staehelina dubia L. a

Family

Ericaceae Cistaceae Cistaceae Cistaceae Thymelaeaceae Leguminosae Leguminosae Cistaceae Globulariaceae Cupressaceae Labiatae Oleaceae Anacardiaceae Fagaceae Fagaceae Rhamnaceae Labiatae Liliaceae Compositae

The study species was only considered in the non-treated experiment.

Species abbreviation

Au Ca Cmo Cs Dg Dh Dp Ft Ga Jo LL Pa PL Qc Qi Ra Ro Sa Sd

Regenerative strategy Seeding

Resprouting

− + + + − + + + + − + − − − − − + − +

+ − − − + + + − + + + + + + + + − + +

Life form

MP NP NP NP NP NP NP C NP MP C NP MP NP MP MP NP PV C

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as mg of each nutrient. P, K, Ca, Mg and Na analyses were undertaken using optical emission spectroscopy with inductively coupled plasma (ICP-OES) in a Perkin Elmer, Optima 4300 (Shelton, Maryland, USA), while N was measured with an elemental analyzer NA 2100 (Thermofisher Sicentific, Milan, Italy). 2.3. Data analysis 2.3.1. Decomposition process To analyze differences in mass loss we considered two variables: decomposition rate (k), and % of remaining mass. Decomposition rate, considered as the annual decomposition constant k, was calculated according to Olson (1963): ln

x  0

xt

= kt

where x0 is the initial litter mass, xt is the mass remaining at time t, and t is the time in years. k Values were estimated for every species and treatments in the climatic experiment, as well as for each species in the non-treated experiment. Differences in k were analyzed for the climatic experiment by a general linear mixed model (GLMM) with climatic treatment (C: control, W: warming, D: drought) and post-fire regenerative strategy (resprouters versus non-resprouters) as fixed factors, and species as a random factor. In the non-treated experiment we also analyzed the effects of postfire regenerative strategy on k-values by means of a GLMM with one fixed factor (regenerative strategy) and species as the random factor. Differences in % of remaining mass were analyzed by repeated measures ANOVA in both experiments. The differences between time (within-subject factor), climate treatments (considered only in the climatic experiment) and post-fire regenerative strategies (between-subject factors) were the factors. To better approximate normality, the % of remaining mass was transformed into its logodd (i.e. log [(x/(1 − x)]). For these statistical analyses, the mean values of each species obtained from the 3 sampled litter bags for each treatment and collection time were used as replicates. We constructed an additional GLMM for the two studied variables (decomposition rate (k) and % of remaining mass (at t = 0.17 years, and t = 2 years)) and also included life-form (fixed factor) and species (random factor) nested within Family. This approach checked that the significance tests for the fixed-effect predictors were not biased by autocorrelations in taxonomical affiliations or by life-form predominance within regenerative strategies. Life-form types were considered as following: C = chamaephyte: with persistent buds situated 0.2–0.5 m height, NP = nano-phanaerophyte: with persistent buds situated 0.5–2 m height, MP = macro-phanaerophyte: with persistent buds situated 0.5–5 m height and PV = phanaerophyte-vine: with persistent buds situated more than 0.5 m height and creeper. Taxonomic order was not considered, as there was nearly the same number of orders as families. 2.3.2. Litter composition First, we analyzed the differences between post-fire regenerative strategies with respect to initial nutrient content, i.e. litter quality, by a one-way ANOVA for each nutrient. The influence of nutrient concentration on the rate of decomposition was analyzed by backward stepwise multiple linear regression. We used k for each species (from non-treated experiment) as the dependent variable and initial N, P, Ca, Na, K, Mg leaf litter concentration and N:P ratio as the independent variables. All variables were transformed to their ln(x). We used a GLMM to test separately whether there was any autocorrelation on account of the higher taxonomical level “family” for each of the nutrients, by considering a hierarchical nested design of

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species (random factor) among families. Order was not considered, as there was nearly the same number of orders as families. This approach ensured that the significance tests for the fixed-effect predictors (climatic treatment and post-fire regenerative strategies) were not biased by taxonomical affiliations. This GLMM also considered life-form as a fixed effect. N, P, Ca, Na, K, Mg concentrations and N:P ratio were log transformed to better approximate normality. To better understand nutrient properties in the different postfire regenerative groups we also analyzed the differences between post-fire regenerative groups for each nutrient concentration in non-treated experiment. For each nutrient, we used a repeatedmeasure ANOVA where time was the within-subject factor (at 0.17, 0.5, 1.17 and 2 years) and post-fire regenerative strategy was the between-subject factor. Analyses were performed using Statistica 6.0 (Statsoft), Sigmaplot 8.0 (SPSS) and SPSS 13.0 (SPSS).

3. Results 3.1. Decomposition process Leaf litter decomposition followed an exponential decay pattern for all species, except for G. alypum and Rhamnus alaternus (see control treatment R2 values in Table C.1), so that these two species were not considered in the statistical analyses of k-values. k-Values and % of remaining mass for all species and treatments are summarized in Table C.1. Drought was the climatic treatment that most affected litter decomposition by decreasing mass loss (Figs. 1 and 2a). Postfire regenerative strategy showed an interaction with time, as resprouters exhibited a lower loss of mass several months after litter decay than non-resprouters (Fig. 1). Significant differences were found in decomposition rates (k) among different climatic treatments (F2,15 = 26.267, p < 0.001). Post-hoc differences indicated that control presented the highest k-values, followed by warming, while drought treatment had the lowest values. There were no differences between post-fire regenerative groups (resprouters, non-resprouters) (F1,15 = 0.028, p = 0.87) or in the interaction between these two factors (F2,15 = 0.684, p = 0.513). When considering life-form type and species nested within families as explicative factors, differences between climate treatments were still evident (F2,15 = 26.833, p < 0.001). There were also significant differences between families (F11,15 = 62.218, p < 0.001), but there were no significant differences between life-forms (F3,15 = 0.057, p = 0.644) or post-fire regenerative strategies (F1,15 = 0.025, p = 0.877). When the decomposition rate for non-treated litter bags (with 4 collection times) was analyzed, non-resprouters presented higher k-values than resprouters (F1,16 = 4.172, p = 0.05). When the taxonomical affiliation and life-form effects were considered, the differences between post-fire regenerative strategies grew weaker (F1,16 = 5.989, p = 0.07) but there were no differences between taxonomical affiliations (F12,16 = 1.926, p = 0.275) or the different life-forms (F3,16 = 0.710, p = 0.595). The analysis of the percentage of the remaining mass in the climatic experiment concurs with the patterns obtained for kvalues, as litter decomposed more slowly under drought conditions (F3,19 = 4.59, p = 0.015), as evident from the higher percentage of remaining mass, while warming had no effect (Fig. 1a). This effect is particularly significant during the first 2 months. After 2 years, the differences between the treatments tend to decrease (Fig. 1a). The percentage of remaining mass did not differ significantly between post-fire regenerative groups (resprouters and nonresprouters) over the 2-year experiment when litter bags from

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(a)

100 Control a Warming a Drought b

aB

Remaining mass (%)

90

aA aA

11

Non-resprouters Resprouters

10 9

80

N content (mg)

(a)

70 bA aA bA

60

8 7 6

B

A

50

5

(b)

Aa

90

4

(b)

Ab

Aa

80

AB AB

BD

CD

0,50

AD

Non-resprouters Resprouters

Ab

0,45 70 Ac Ac

Ad

60

Ac A

50 0,0

B

0,5

C

1,0

D

1,5

2,0

2,5

P content (mg)

Remaining mass (%)

Non-resprouters(S+ R-) Resprouters (S- R+, S+ R+)

Time (years)

the climatic treatments experiment were considered, but there was a significant interaction between time and post-fire regenerative strategy (F3,19 = 4.19, p = 0.034). Also, when the percentage of remaining mass was analyzed over time in non-treated data, there was a significant interaction between regenerative strategy and time (F3,19 = 4.19, p = 0.001), meaning that non-resprouters have a faster decomposition than resprouters (Fig. 1b). The interaction between climatic treatment and post-fire regenerative strategy was not significant, so that climatic treatment did not affect post-fire regenerative litter decomposition patterns. Life-form and taxonomical affiliations showed that, apart from the differences between treatments and post-fire regenerative groups, there were also differences between taxonomical affiliations, but not between life-form types (Table 2). 3.2. Litter composition The results of the chemical analyses of the initial litter samples are shown in Table C.1. There were no significant differences between post-fire regenerative strategies in relation to the initial litter quality (N: F1,19 = 0.018, p = 0.894; P: F1,19 = 0.475, p = 0.499; Ca: F1,19 = 2.193, p = 0.157; K: F1,19 = 0.016, p = 0.900; Na: F1,19 = 0.074, p = 0.789; Mg: F1,19 = 0.276, p = 0.606; N:P: F1,19 = 2.905, p = 0.107). The results of the multiple regression analyses of k (from non-treated experiment samples) in relation to initial chemical

0,35 0,30 0,25 0,20 0,15

(c) N : P content ratio

Fig. 1. (a) Mean of the remaining mass (%) for each climatic treatment in the two collection times. (b) Mean of the remaining mass (%) for each post-fire regenerative strategy in the non-treated experiment (with 4 collection times). Vertical bars denote standard errors. Post-hoc Bonferroni’s significant differences (p < 0.05) are indicated with different letters. Upper case letters (A and B) indicate differences between treatments (a) or between regenerative strategies (b) for a given collection time. Lower case letters (a and b) indicate differences between collection times for a given climatic treatment (a) or between post-fire regenerative strategies (b). Italic upper case letters (A and B) indicate differences between collection times, italic lower case letters (a and b) in the legend box indicate differences between treatments (a).

0,40

40

A A

B

C

C

B

C

Non-resprouters a Resprouters b

35 30 25 20 15

A A 0,0

A 0,5

1,0

1,5

2,0

Time (years) Fig. 2. Mean of the N, P and N:P litter content for each regenerative strategy in the four collection times considered outside the climatic treatments. Vertical bars denote standard errors. Post-hoc Bonferroni’s significant differences (p < 0.05) are indicated with different letters. Italic upper case letters (A and B) indicate differences over time. Italic lower case letters (a and b) indicate differences between regenerative strategies. There were non-significant differences between collection times for a given regenerative strategy.

concentration suggest that 48% of the variance (R2 ) (F = 4.546, p = 0.019) was basically explained by only a few of the litter composition variables. The nutrients that best fitted the model were: nitrogen (t = 3.154, p = 0.007), calcium (t = 1.961, p = 0.069) and sodium (t = 1.961, p = 0.139) (leaf litter concentration = −0.208 + 0.172 (ln(N)) + 0.007 (Ca) + 0.119 (Na)). When analyzing each nutrient separately after 2 months and 2 years of decomposition (Table 3), nitrogen (time 0.17 years:

S. Saura-Mas et al. / Environmental and Experimental Botany 77 (2012) 274–282 Table 2 General linear mixed model (GLMM) for the remaining mass (%) at different collection times (0.17 and 2 years). (a) Climatic treatment and regenerative strategy were the fixed factors and species was the random factor. (b) The same GLMM with two more factors, life-form as another fixed factor and species (as a random factor) nested among families. Life-form types and families were considered as in Table 1.

(a) Climatic treatment Regenerative strategy Species (b) Life-form Climatic treatment Regenerative strategy Family (species) a *

Remaining mass (t = 0.17)

Remaining mass (t = 2)

F

p

F

p

84.340 0.713 18.351

<0.001* 0.412 <0.001*

21.481 0.010 57.685

<0.001* 0.922 <0.001*

2.831 84.340 0.973 13.433

0.083 <0.001* , a 0.343 <0.001*

1.345 21.481 0.011 53.958

0.306 <0.001* , a 0.920 <0.001*

279

4. Discussion Climate manipulations modified the pattern of litter decomposition of the set of studied Mediterranean species. Experimentally induced drought diminished the loss of litter mass over time, leading to higher values of remaining litter after 2 years of experiment. This pattern may be explained by climatically controlled modifications to the structure and activity of the microbial population and the humification process (Coûteaux et al., 1995). Accordingly, reduced microbial and fungus activity under drought conditions results in a lower decomposition rate and lower mass loss, so that nutrient cycling and functions of the ecosystem may be altered by a reduced incorporation and recycling of organic matter into the soil. Soil humidity has been proposed as the key factor affecting decomposition rates in Mediterranean ecosystems (Cortez, 1998) ˜ et al., 2006) and, more specifically, it has been suggested (Ormeno that soil moisture greatly improves leaf litter colonization by fungi, responsible for exerting the greatest influence on the temporal and spatial variability of soil organic-matter decomposition and recycling (Curiel-Yuste et al., 2011; Butenschoen et al., 2011). The differences between climatic treatments were more evident in the first months of decomposition. This is probably because losses in this phase are mainly derived easily degraded soluble compounds and celluloses. The results suggest that these soluble compounds may lixiviate with more difficulty under drought treatment. After 2 months, the remaining mass is mainly composed of lignified material which is not as easily degraded (Coûteaux et al., 1995). Warming did not affect decomposition in our experiment, even though a stimulatory effect of temperature on the decomposition rate could be expected. Temperature is not likely to be a limiting factor for decomposition in Mediterranean ecosystems, and the relatively small increase in temperature (0.6–0.7 ◦ C) induced by the experiment may have been insufficient to produce this stimulatory effect, or may have been counterbalanced by the associated increased in aridification. Thus, this study suggests that under the current trends of climatic change, litter decomposition in Mediterranean ecosystems would probably be more diminished by water depletion than increased by a moderate rise in temperature. More studies on the synergies between temperature increase and drought need to be conducted, however, to better understand the effects of climate change. Litter decomposition varied between species belonging to the different post-fire regeneration groups. The physical attributes of leaves can be one of the causes of these differences. Recent studies (Garnier et al., 2001) have shown differences between species leaf dry matter content (LDMC) in saturated conditions (LDMCsat ) and. More specifically, Saura-Mas and Lloret (2007) have demonstrated that resprouters have higher LDMCsat than other post-fire regenerative groups. A high LDMCsat corresponds to a low proportion of mesophyll and epidermis (light tissues) and a high

Post-hoc Bonferroni differences were found in drought treatment. Significant results (p < 0.05).

F = 7.475, p = 0.002; time 2 years: F = 14.577, p < 0.001) and N:P were significantly lower in drought treatment (time 2 years: F = 4.406, p = 0.02), while K was higher in this treatment (time 0.17 years: F = 85.94, p < 0.001). Other differences between climatic treatments were found in Ca (time 2 years: F = 3.86, p = 0.031), Na (time 0.17 years: F = 5.27, p = 0.011) and Mg (time 0.17 years: F = 24.649; p < 0.001) (Table 3). Significant differences were found between life forms. Macro-phanerophytes presented the lowest Na and Mg concentrations and nano-phanerophytes and chamaephytes presented the highest values. Resprouters presented lower Mg and K concentrations than non-resprouters at time 0.17 years. Finally, taxonomical affiliation exerted an effect on all the concentrations of nutrients (Table 3). We also analyzed all the leaf nutrient contents in the nontreated experiment. All groups of species tended to reduce their litter nutrient contents over time (p < 0.05), especially with respect to N, P, Na and K (Fig. 2). The content of some nutrients, such as Mg and Ca, decreased at first but after a while it tended to increase again (Fig. D.1). In general, non-resprouters tend to lose less nutrients but only the N:P ratio of litter is significantly lower in non-resprouters than in resprouters (F4,19 = 2.842, p = 0.030, Fig. 2c), indicating that the litter of non-resprouters has a higher proportion of P than resprouters. The N:P ratio increased over time, indicating that in both groups P return to the soil is more rapid than N return. This is in keeping with the results for N content and P content, as nonresprouters tend to have lower values of N and higher values of P than resprouters (Fig. 2a and b). The litter Mg concentration shows a significant interaction between post-fire regenerative strategy and time (F1,19 = 2.59, p = 0.009), so non-resprouters tend to lose less Mg over time than resprouters (Fig. D.1).

Table 3 Mean and standard error of litter nutrient concentrations (mg g−1 ) and N:P ratio (for the two collection times) for each climatic treatment and regenerative strategy. Lower case letters indicate Bonferroni post-hoc differences among climatic treatments or post-fire regenerative strategies from significant results of GLMM. Time (year)

N (mg g−1 )

P (mg g−1 )

Ca (mg g−1 )

K (mg g−1 )

Na (mg g−1 )

Mg (mg g−1 )

N:P Mean

Mean

S. E.

Mean

S. E.

Mean

S. E.

Mean

S. E.

Mean

S. E.

Mean

S. E.

Control Drought Warming Control Drought Warming

0.17 0.17 0.17 2 2 2

10.167b 9.585a 10.291b 11.801 11.097 11.527

1.065 1.089 1.109 1.153 1.164 1.182

0.392 0.372 0.389 0.397 0.387 0.394

0.054 0.055 0.057 0.054 0.051 0.056

21.707 21.555 21.311 37.217ab 37.875a 35.112b

1.835 1.684 1.692 2.956 3.132 2.968

1.893b 4.489a 2.278b 0.986 0.919 0.939

0.274 0.627 0.372 0.122 0.103 0.126

1.158a 1.187ab 0.958b 0.145 0.136 0.145

0.129 0.185 0.170 0.017 0.018 0.022

0.288b 0.743a 0.563b 1.170 1.235 1.238

0.083 0.183 0.184 0.115 0.122 0.130

Resprouters Non-resprouters Resprouters Non-resprouters

0.17 0.17 2 2

10.350 8.924 11.607 11.046

0.742 1.015 0.765 1.359

0.372 0.426 0.369 0.470

0.038 0.048 0.036 0.051

21.092 22.927 36.164 38.590

1.190 1.598 2.050 3.050

2.874 2.927 0.804 1.415

0.353a 0.600b 0.058 0.150

1.211 0.742 0.125 0.198

0.101 0.197 0.009 0.031

0.387 1.000 1.133 1.480

0.098a 0.179b 0.078 0.126

28.339 28.669 29.690 32.826a 31.043b 32.126ab 308.016 227.586 345.217 237.975

S. E. 1.905 2.282 2.288 2.413 2.279 2.390 13.288 21.898 15.007 11.847

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proportion of vascular tissues and sclerenchyma (dense tissues) (Dijkstra and Lambers, 1989; Niemann et al., 1992; Garnier and Laurent, 1994). Kazakou et al. (2006) found a strong negative correlation between LDMCsat and decomposition rate. Accordingly, resprouters are expected to present higher litter mass after 2 years of experiment. This prediction is supported by our results (Fig. 1b), since the loss of litter mass of resprouters observed 2 months after the start of the experiment was lower than in non-resprouters, and these differences had increased at the end of the experiment. Our results show that differences in decomposition rate and litter chemical composition between climatic treatments are influenced by the taxonomical affiliation of the studied species. Taxonomical affiliation plays an important role in determining the decomposition rate, probably because the physical properties of the leaf that determine decomposition rate, such as LDMCsat , are determined phylogenetically (Saura-Mas and Lloret, 2007). Many non-resprouters (seeders) evolved during the Quaternary, when Mediterranean conditions were established, within a relatively low number of families, while many resprouters were already present in the Tertiary in a wide range of families (Herrera, 1992; Verdú, 2000; Pausas and Verdú, 2005). Therefore, the life strategy and leaf attributes of seeders has been postulated as an adaptive response to the Mediterranean climate, and this concurs with studies that show evidence of seeders’ advantages for surviving under climates ˜ with severe drought events (Penuelas et al., 2001; Lloret et al., 2005; Arnan et al., 2007). The chemical and physical characteristics of litter material regulate the decomposition rate (Gallardo, 2001). Coûteaux et al. (1995) suggested that climate would regulate the decomposition rate in ecosystems under unfavorable climates, while in ecosystems with favorable climates the chemical composition of the leaf litter would be the best predictor of decomposition. The main components of the leaf litter from senescent leaves of woody species are celluloses, hemicelluloses, lignines and lipids (in order of abundance), and each of these components decomposes at a different rate (Gallardo, 2001). Furthermore, there is a positive relationship between the microbial activity and the nutrient abundance of the leaf litter limiting microbial productivity (Gallardo, 2001). Our results confirm these assumptions, as species with higher initial nitrogen concentration show higher k-values. Our study suggests that Ca and Na concentrations could also play an important role in decomposition microbial activity, at least in soils with similar characteristics (petrocalcic calcixerept). Most decomposition studies have found increases in the relative concentration of nutrients over time (Kavvadias et al., 2001; Allison and Vitousek, 2004; Hirobe et al., 2004; Maisto et al., 2011). Translocation of nutrients from soil to leaf litter via fungus hyphae have been proposed as an explanation for N increases, and some calculations indicate that 35% of the N incorporated into leaf litter was derived from fungus nitrogen (Gallardo and Merino, 1992; Zeller et al., 2000). In fact, our results show a trend toward maintenance of the levels of nitrogen, albeit with some differences from one treatment to another. In fact, nitrogen deposition may also be depleted through rain interception in the drought treatment, which could explain the lower values of litter nitrogen concentration. As leaf litter concentration of nutrients is dependent on the remaining mass, this study also considered the absolute values of nutrient contents. We found that most of the litter nutrients contents tend to decay over time, which was to be expected, on account of degradation, lixiviation and other decomposition processes. K and Na return to the soil rapidly, probably because of their facility for lixiviation. This corresponds with the finding that litter was richer in K concentrations under drought treatment. Ca and Mg do not seem to follow this pattern of rapid return to the soil. This study suggests that, in general, non-resprouters show higher nutrient concentrations in the litter fraction, compared

to resprouters. This trend could be explained by the fact that non-resprouters present greater more mass losses, so that the concentration of nutrient per leaf mass is likely to increase. The main difference between post-fire regenerative strategies is that nonresprouters present greater Mg loss and tend to have a lower N:P ratio. Thus, ecosystems dominated by non-resprouters are expected to have a higher proportion of P in the leaf litter fraction, in relation to N. This is probably because they do not accumulate as much N in the leaf litter as resprouters and because P concentration is higher in live leaves from non-resprouter species (Saura-Mas and Lloret, 2009a). Lower Mg loss in non-resprouters means that communities dominated by this group of species will have slower Mg fluxes, which could be related to photosynthesis and other cell properties. The different role of the two post-fire regenerative types in nutrient recycling may have important consequences at community level when changes in the disturbance regime are considered. The respective prevalence of resprouters and non-resprouters as a result of post-fire regenerative strategies may be determined in the future by new fire regimes driven by climate change; this is ˜ et al., 1998; Pausas, likely to increase the occurrence of fires (Pinol 2004) and may therefore alter the balance between these two types of plants. Vila-Cabrera et al. (2008) have anticipated an increase in short-lived non-resprouter species with increasing fire frequencies, although a very high rate of recurring fires is likely to result in communities dominated by herbaceous species. Concluding, our results experimentally highlight the potential effects of climate change on a key ecosystem process as litter decomposition, and these effects interact with species attributes that respond to a complex network of environmental factors, particularly disturbance regime, such as wildfires. Thus, ecosystems dominated by resprouters in the Mediterranean basin tend toward slower decomposition over longer periods of time in comparison with communities dominated by non-resprouters. Moreover, the leaf litter of the former communities will show a higher proportion of N in relation to P. These changes in the flux of matter and nutrient, combined with a potential shift in community composition, would complement other changes resulting from the predicted increase in drought periods induced by climate change, as observed in experimental climatic manipulation. Acknowledgments We would like to thank M. Cabezas, J. Mas, M.T. Mas, and A. Saperas for helping in the laboratory and A. Vila and M. Sallent for helping in the field. Special thanks are due to D. Sol and B. Shipley for their support and suggestions. This study was funded by the Department of Universities, Research and Information Society of the Generalitat de Catalunya (2009-SGR-247 and 2009-SGR-458), the European social funds, and the Spanish MCYT projects REN 2003-07198, CGL 2006-01293/BOS, CSD 2008-00040, CGL 20090810 and CGL 2010-17172. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.envexpbot.2011.11.014. References Ackerly, D.D., 2004. Functional strategies of chaparral shrubs in relation to seasonal water deficit and disturbance. Ecological Monographs 74, 25–44. Alberdi, L., Cavero, R., 2003. Flora vascular post-incendio en un carrascal de Nazar (Navarra). Publicaciones de Biología, Universidad de Navarra, Serie Botánica 15, 1–17. Allison, S.D., Vitousek, P.M., 2004. Rapid nutrient cycling in leaf litter from invasive plants in Hawaii. Oecologia 141, 612–619.

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