Author: Halabuk, Andrej
Date published: January 1, 2011
Invasive species influence ecosystem processes in a complex way. According to many examples in literature, an invasive species alters processes and subsequent species composition and biodiversity in the ecosystem (Levine and D'Antonio, 1999; Levine et al., 2003; ViTousEK, 1990).
When the species composition has been changed there is a high probability that the subsequent soil-plant interactions and carbon and nutrient cycles will change, too. These consequences may, in turn, have an impact on the invasive susceptibility of ecosystems and the invasivness of species (Ehrenfeld, 2003). Invasive and expansive species have a potential to change many components of the carbon and nitrogen cycles of an ecosystem. Ehrenfeld (2003) has summarized, that invasive plant species frequently increase phytomass production, net primary production, nitrogen availability, alter nitrogen fixation rates and produce litter with higher decomposition rates than the co-occurring native species.
Leaf litter decomposition plays an important role in caibon and nitrogen cycling in ecosystems, being a main source of nutrients and organic matter for plants and soil organisms. Leaf litter decomposition is affected by both biotic and abiotic factors: temperature, moisture, litter chemistry, soil nutrient supply and decomposer community structure. Due to the large number of factors controlling the leaf litter decomposition, comprehensive study of impact of invasive species on leaf litter decomposition rate is difficult to pursue. The impotance of specific factors and their interactions are often studied in order to find the relative effect of each particular factor. Furthermore, specific habitat conditions can affect microbial activity and leaf litter decomposition significantly (Simonovicová, 1995; Lancuch and Simonovicová, 2008). Comparative studies of invasive species and co-occurring native species in field conditions can provide valuable information on this topic.
Disturbed ecosystems are examples exhibiting obvious invasions of exotic and expansive plant species. Clear-cut areas and forest openings with high nutrient availability in soil are veiy sensitive to invaders. The invaders are mostly very effective in nitrogen using. Under conditions of forest opening they can invade extremely (Elias, 2000). Even though it is probable that exotic plant invasions may alter soil and ecosystem properties, it is not always the case (Ehrenfeld and Scott, 2001). In addition, environmental conditions much varying on a grathent from clear-cut to closed forest gradient may result in a very variable decomposition rate (DiDHAM, 1998). Therefore, there is still a need for case studies on different invasive plant species in various types of habitats.
In our contribution, we compare the decomposition rates of selected exotic species with relevant native species. We select Robinia pseudoacacia and Ailanthus altissima from the exotic expansive woody species and Impatiens parviflora and AlHaHa petiolata from expansive herbaceous species. Although Alliaria petiolata is not an exotic species, it is considered as native expansive species. Acer campestre and Mercuri ali s perennis were selected as native co-occurring species for comparative analysis. There are many studies on ecology of these invasive species in Slovakia; to our knowledge, however, there has not been performed yet a study of litter decomposition rates in these species in Slovakia.
Material and methods
The research site is located in the SW Slovakia (48°18?9"; 17°53'27 ") in the Báb forest situated in a warm and dry region, with the mean annual temperature of 9.3 0C and precipitation total of 580 mm (Tuzinsky, 2004). The parent material is calcareous loess, the formed soils are Albi-Haplic Luvisols with 2.6-3.3% organic matter content and C/N ratio of 13.9 in the 0-10 cm horizon (Sombathová and Zaujec, 2001; Gonet et al., 2008). The geobiocoenosis belongs to the FagetoQuercetum group of forest type (Kubícek and Brechtl, 1970), with prevailing oaks - Quercus robur, Q. petraea, Q. cerris with admixture of Carpinus be tu lus and Acer campestre in the tree layer of phytocoenosis. The herbaceous layer is dominated by Mellica uniflora, Dactylis glomerata, Carex pilosa, Asperula odorata, Hederá helix, Pulmonaria officinalis and Geum urbanum. Since 1966, a part of the Bab forest is protected. In this part has been prohibited forest management. In the adjacent area, 4 clear cuttings were applied at the end of the year 2006. The clear-cut areas were consecutively massively occupied by Ailanthus altissima, Robinia pseudoacacia and A lliaria petiolata species.
Litter bag experiment
For determination of leaf litter decomposition, we used the standard litter bag method (see Harmon et al., 1999). Leaf litter was collected in late summer when senescence was evident. In case of tree species, the litter was collected from branches immediately before shading of leaves in October 2008. The collected material was air dried in the laboratory for 14 days and filled in bags in amounts of 3-5 g. The initial leaf moisture content was estimated with subsamples set aside from litter samples and dried at 75 0C. Our litterbags were 16 cm ? 12 cm in size, made of nylon mesh (1 mm mesh size), with three sides double-stitched with nylon thread, and provided with small plastic tags with identification numbers. We established five plots on a grathent from clears-cut area to closed forest (Fig. 1). The litter bags were placed randomly on particular plots, directly on the ground under the recent litter. The number of litterbags of individual species on particular plot varied from 30 to 50. Altogether we used 500 litterbags in the field. Together 4-6 samples of every species were taken in field in each sampling period. The sampling dates for heibaceous litter were: 4 September 2008, 17 October 2008, 14 November 2008, 19 December 2008, 27 January 2009, 19 March 2009 and 24 My 2009; for woody species: 17 October 2008, 19 December 2008, 27 January 2009, 19 March 2009, 24 July 2009 and 25 September 2009. The species determination was made according to Dostál and Cervenka (1991) and the nomenclature follows that of Marhold and Hindák (1998).
The standard cellulose decomposition experiment was made with the aid of cellulose filter paper inserted in the litter bags (Skolek, 1980). The litter bags with cellulose filter paper were placed under the leaf litter (ca 1 cm in the topsoil) and exposed for 2 months in the field in spring 2009. Five replicates were used on each plot, together 25 litter bags. Then the litterbags were transported in the laboratory, cleaned of adhering waste (soil, mosses, rock fragments, etc.), oven dried at 75 0C until a stable weight (usually ca. 48 hours), and weighed to determine the dry weight. Two-way anova (StatSoft, INC., 2010) was used in order to reveal significant differences for testing differences in litter decomposition rate among different species across the study site. Herbaceous and tree species were analysed separately. The annual profile of decomposition was analyzed by 1 -phase and 2-phase negative exponential decay model (Olson, 1963; Lindsay and French, 2004). The K decomposition constant, half-lives and 90% decay life and R coefficient of determination of the fitted model were obtained by means of GraphPad Prism software (GraphPad Software, 2009).
The near-ground (30 cm) air temperature was monitored in clear cut area and closed forest canopy with HOBO Pro temperature loggers (ONSET Computers, USA) at 30 min intervals. There was analysed the period from 4 September 2008 to 21 September 2009. Volumetric soil moisture was monitored in 30 min interval with Virrib sensors (AMET, CR) based on phase transmittance method at 10, 30 and 70 cm depth in clear-cut area and closed forest.
The decomposition rate leaf litter in herbaceous species varied substantially both among herb species and in time (Fig. 2). The biggest differences in weight of herb species litter are visible before the winter, after exposition in the field one month (1st sampling date) where all mean differences in litter weight were even statistically significant (p < 0.05). Later, after 100 days in the field, differences in weight of herb species litter continuously decreased (Fig. 3), mainly in case oí Alliaria petiolata and Mercurialis perennis, where also the inverse pattern occurred on 4th and 5th sampling date. After 324 days of exposure in the field, conspicuous loss of litter weight was detected in case oî Impatiens parviflora species (74%), followed by Alliaria petiolata (71%) and Mercurialis perennis species (71%). However, only the mean difference in litter weight loss between Impatiens parviflora and Mercurialis perennis species was statistically significant. Mean differences in litter loss of all studied species and their statistical significance for every sampling date are presented in Table 1.
In case of woody species, the litter loss differences were statistically significant during the whole monitoring period. The differences slightly increased during the time period in the field. Especiall, the low decomposition of Robinia pseudoacacia litter in the later period is noticeable (Fig. 4). After 344 days of exposure in the field, the great loss of litter was detected in case of Ailanthus altissima species (55%), followed by Acer campestre (37%) and Robinia pseudoacacia species (30%).
The values of regression coefficients show the two-phase exponential decay model as better explaining the litter decomposition in all the plant species than the one-phase exponential model (Table 2). However, the results of the one-phase exponential decay model allow standard comparison with other published results. The first phase of litter decomposition (50% of weight) was found the fastest in Impatiens parviflora species, followed by Alliaria petiolata and Mercurialis perennis. The substantial decrease of decomposition is evident in winter season; however, in later periods the decay profile is very similar for all species (Fig. 5).
The rate of cellulose decomposition varied significantly among the habitats, with the maximum in the closed forest (Fig. 6). This pattern was to extent similar to the decomposition pattern in herbaceous species litter (Fig. 7). However, the results of anova proved the habitat effect to be significant only for the species Mercurialis perennis and Impatiens parviflora. This habitat effect was not observable in woody species (Fig. 8). However, the evident considerable variability in these values may be caused by many factors of habitat affecting the decomposition rate of litter. Varying site microclimate was detected throughout the forest clear-cut area to closed forest grathent (Fig. 1, Fig. 9). The mean difference in values of near-ground air temperature recorded in clear-cut area and in closed forest during the period from 4 September 2008 to 21 September 2009 made 0.36 0C. However, the mean daily temperature difference (7.00 a.m.-18.00 p.m.) made 2.76 0C. In daily profile, the greatest differences were identified between 13.00 and 14.00 p.m. Similarly, the values of mean difference of relative air humidity recorded during the period from 22 April 2009 to 2 1 September 2009 in clear-cut area and closed forest made 12% and the mean daily difference 2.3% (Fig. 10). The soil moisture value in the clear-cut area was permanently higher compared to the soil moisture in the closed forest (Halabuk, 2010). The biggest differences in soil moisture values recorded in the clear-cut area and closed forest were detected mainly at a depth of 70 cm. In topsoil, where cellulose decomposition was measured, the moisture differences were not such conspicuous, probably due to the higher input of precipitation water and due to dew formation.
It is obvious that decomposition rate of herbaceous litter is faster than that of tree leaf litter (Mayer, 2008). Decomposition of plant litter is strongly affected by plant litter quality. The rate of microbial decomposition of plant litter is usually highly positively correlated with increasing N concentration and negatively correlated with increasing C/N ratios in this litter, probably reflecting differences in its structural and secondary compounds, such as lignin and phenols (Túma. 2002). This variable chemistry of litter during decomposition results in a two stage decay profile exhibiting evidently faster decomposition of labile compounds at the beginning of exposure in the field. That is why the two-phase decomposition model usually better explains decomposition within a shorter period (Lindsay, 2004); and allows, in such a way, better comparison of results of decomposition of different species litter (Didham, 1998). However, better long-term prediction of litter decomposition pattern requires longer exposition of samples in the field (e.g. during the next winter), which in turn would need more litter of herbaceous species in the bags. On the other hand, regarding the studied effect of invasive plant species on mineral cycling during decomposition, the short term decomposition, with the major part of biomass decomposed is important. Invasive species considered having higher leaf N content and lower C/N ratios, usually decompose faster (Ehrenfeld, 2003; Ashton, 2005; Lindsay, 2004). The fast decomposition rate of rich-in-nutrients leaves of Ailanthus altissima (Castro-Diez, 2009), Robinia pseudoacacia (Tateno et al., 2007), Alliaria petiolata (Rodgers, 2008) and Impatiens parviflora species (Vanderhoeven et al., 2007) has been documented; although their decomposition differences have never been compared in the same environmental conditions. Our results showed that the litter decomposition of invasive species was faster compared to the indigenous species, with exception of Robinia pseudoacacia, leaf litter. This was a surprise as Robinia pseudoacacia is a nitrogen fixing species with high N content and low C/N ratio in leaves. This possible explanation is high lignin content in locust leaves increasing their resistance against most decomposing organisms - as it has been also documented by CastroDiEs (2009).
Despite the well documented fact that the leaf litter quality of plant species is a main driver of differences in rate of its decomposition (Funk, 2005; Standish et al., 2004), the habitat (site) effect should also be considered in order to identify additional factors possibly affecting the decomposition. Furthermore, canopy of invasive tree species may create a specific environment facilitating decomposition of plant litter irrespective of its quality (Ashton, 2005), providing, in such a way, positive feedback for expansion of this species.
Variable soil moisture and temperature across a disturbed forest could affect plant decomposition (Tesaro va, 1993). Logging of forest stands is accompanied by extensive organic matter decomposition and humus mineralization, since conspicuous changes in microclimate on clear-cut areas create more favourable conditions for the activity of soil micro-organisms (Zahora, 1996; Tuma, 1998). However, Tűma (2002) ascertained only positive effect of temperature on decomposition rate. Negative correlation between soil moisture and microbial decomposition rate may reflect possible slower heating of waterlogged soils on clear-cut areas (Tuma, 1999). In our study, there was a consistent effect of plant species leaf litter across the site, since in the two-way Anova, there was no significant interaction between the site effect and the effect of plant species leaf litter. However, we did not register a higher decomposition rates in clear-cut area. Furthermore, herbaceous species litter and cellulose tests proved higher decomposition in the closed forest compared to clearcut area. We documented that clear-cut area was wetter and warmer than the closed forest, however these differences vary in dependence of the season. Bublinec (1975) stated that the greatest decomposition rates were evident at the beginning of vegetation season and they substantially decreased in August. The author also pointed to the soil moisture as possible cause of temporal dynamics of decomposition. The greatest differences in microclimate between closed forest and clear-cut area, however, where ascertained in summer, being low in autumn and early spring.
In a well designed multifactor experiment by Mayer (2008), the author documented that decomposition rate is more affected by litter moisture than soil temperature, which can explain the higher decomposition loss of plant litter under forest canopies. In fact, higher nearground air humidity in a closed forest during day time (Fig. 10) may contribute to surface wetness in the litter horizon. Furthermore, litter decomposition in old forest stands is also influenced by communities of decomposers adapted to consuming the large stocks of leaf litter (Mayer, 2008). The specific differences in composition of soil micro-organisms and lower microbial activity in Báb clear-cut area compared to the adjacent closed forest may refer to this fact (Dugová, 2008). We also need to note that by using mesh litter bags in our study we excluded a well known effect of macrodetrivores on decomposition rate (Mayer, 2008), which may underestimate the faster decomposition of plant litter in well structured forest. However, in the near future, massive expansion of invasive species across the clear-cut area may change this pattern, as these species are able to change the microclimate (Standish, 2004) and community of decomposers at the site (Holly, 2009. The issue requires several years of additional experiments.
We can summarise that, except Robinia pseudoacacia, leaves of exotic and native invasive and expansive species were decomposing faster, mainly in the first stage - decomposition of the most valuable litter parts with high nitrogen content. This pattern was the same across the closed forest, forest edge and clearing. Although the clearing was wetter and warmer on average, decomposition rates of cellulose and of herbaceous litter were higher in the closed forest.
This research was supported by the Slovak Grant Agency for Science VEGA, projects No 2/7132/07 and No 2/0174/10. We thank the editor and to anonymous reviewers for their comments and suggestions on the earlier version of the manuscript.
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Andrej Halabuk, Katarina Gerhátová
Institute of Landscape Ecology of the Slovak Academy of Sciences, Branch Nitra, Akademická 2, 949 01 Nitra,