Effect of N addition on home-field advantage of litter decomposition in subtropical forests

Forest Ecology and Management 398 (2017) 216–225

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Effect of N addition on home-field advantage of litter decomposition in subtropical forests Hong Lin, Zaihua He, Jiewei Hao, Kai Tian, Xiuqin Jia, Xiangshi Kong, Siddiq Akbar, Zhanlin Bei, Xingjun Tian ⇑ School of Life Sciences, Nanjing University, Nanjing, China

a r t i c l e

i n f o

Article history: Received 9 February 2017 Received in revised form 7 May 2017 Accepted 8 May 2017

Keywords: Home-field advantage Litter decomposition Subtropical forests Specialization Nitrogen cycling

a b s t r a c t The ‘home-field advantage’ (HFA) postulates that litter decompose faster in its home habitat than in other habitats. However, the HFA of litter decomposition appears to be highly variable, and the effects of environmental conditions on HFA have rarely been investigated. Thus, in this study, we performed a reciprocal litter transplant experiment using coarse and fine mesh litterbags under nitrogen (N) addition treatments in a subtropical coniferous forest dominated by Pinus massoniana and a broad-leaved forest dominated by Quercus variabilis. Results showed no significant difference in decomposition between the two dominant litters in the fine-mesh litterbags at home and away habitats under control and N addition plots. P. massoniana litter in the coarse-mesh litterbags decomposed twice as fast at home than in away habitats under N addition. The result suggests a positive HFA effect in the coniferous forest under N addition, but no significant HFA effect was observed in the control plots. N addition did not enhance Q. variabilis litter decomposition in the home habitat. The positive HFA effect of P. massoniana litter in the coarse-mesh litterbags in N addition plots was associated with more abundant soil fauna than in the control plots. However, N addition had no significant effect on the activity of most soil enzyme during litter decomposition. Moreover, soil microbial biomass showed no relationship with the HFA of litter decomposition. Our findings suggest that N addition likely enhances the feeding activity of soil fauna by increasing fauna abundance. This further reinforces the habitat specificity of soil mesofauna in coniferous forests, resulting in a positive HFA of P. massoniana litter decomposition. C and N cycling in coniferous forest may be enhanced by N addition, and coniferous forest management should be suitable for this change. Ó 2017 Elsevier B.V. All rights reserved.

1. Introduction Litter decomposition is a fundamental process in the functioning of forest ecosystems as it facilitates the recycling of nutrients and chemical elements, and regulates forest restoration and productivity (Cleveland et al., 2011). Physical and chemical environment (e.g., temperature, humidity, and UV), litter quality (e.g., C: N, lignin:N), and soil decomposers (e.g., bacteria, fungi, and invertebrates) affect decomposition (Hättenschwiler et al., 2005; Prescott, 2005). Different plant litters have different physical and chemical properties, and different organisms are involved in breaking down the materials, which affect the rate of decomposition. Decomposers may particularly decompose litter from a specific plant with which they are associated (Veen et al., 2015a). Previous research suggests that litter may decompose faster in the habitat ⇑ Corresponding author. E-mail address: [email protected] (X. Tian). http://dx.doi.org/10.1016/j.foreco.2017.05.015 0378-1127/Ó 2017 Elsevier B.V. All rights reserved.

from which it was derived than in other habitats. This phenomenon has been called the ‘home-field advantage’ (HFA) effect (Gholz et al., 2000; Ayres et al., 2009). After long-term competitive adaptation, soil decomposers and litter form coexistence mechanisms because litter is the main source of nutrients and energy for soil organisms. There is evidence that demonstrate the importance of substrate–microbial interactions leading to HFA (Strickland et al., 2009a). Plant species can influence the activity of the soil microorganism community directly through leaching or release of exudates (Pfeiffer et al., 2013), or indirectly by affecting competitive interactions among soil decomposers (Cesarz et al., 2013; Austin et al., 2014). With the continual input of a similar litter quality, microbial decomposers prefer to decompose this particular kind of litter (Ayres et al., 2009), which could generate specificity of decomposers for a particular plant species and its associated habitat. Ayres et al. (2009) calculated the magnitude of HFA in forest ecosystems, showing that positive HFA accelerated litter mass loss

H. Lin et al. / Forest Ecology and Management 398 (2017) 216–225

by approximately 8% using reciprocal transplant experiments. However, Gießelmann et al. (2011) showed that both microorganisms and mesofauna had no significant effects on HFA of litter decomposition in an Atlantic rainforest. Similarly, St. John et al. (2011) found no HFA effect in a forest–grassland reciprocal transplant experiment and attributed the result to an adaptation of soil microbial communities to different litter resources. These studies suggest that HFA, based on specialization of soil microbial communities, does not exist and the magnitude and direction of the HFA effect varies considerably (Ayres et al., 2009; Veen et al., 2015b). Differences in leaf litter quality could explain the occurrence of HFA. Low-quality litter that contains recalcitrant or toxic secondary compounds may generate a large HFA because fewer soil communities can decompose these compounds (Austin et al., 2014; Chomel et al., 2015). By contrast, high-quality litter, which contains easily degradable compounds, could be expected to have a lower HFA because most soil decomposers can decompose them (Ayres et al., 2009; Austin et al., 2014; Veen et al., 2015b). The litter quality of subtropical coniferous forests and broad-leaved forests differ largely in terms of initial N, P, and lignin concentrations, C: N and lignin:N ratios, as well as secondary compounds (Ushio et al., 2012; Wang et al., 2012). The effects of microbial specialization on litter decomposition can be demonstrated by extracellular enzyme activity (Wallenstein et al., 2013; Chomel et al., 2015). Microbes produce specialized enzymes to break down complex compounds found in recalcitrant litter. Bacteria, such as Actinomycete sp., release peroxidase, esterase, and oxidase to decay humus and lignin. Fungi, such as Trichoderma sp. and Pythium sp., release acids and alkaline phosphatase (ALP), urease (URE), b-glucanase, cellulase, and chitinase (Naseby et al., 2000; Zhang et al., 2004). Extracellular enzymes perform different activities according to the litter matrix composition. For example, in the forests where litter has high contents of lignin, the biomass of lignin-degrading fungi tends to be more abundant than other fungi. Recalcitrant and low-quality litter requires more specialized enzymes to break down these compounds (Chomel et al., 2015). In contrast to higher-quality litter, low-quality litter is more likely to generate an HFA effect due to specialization of extracellular enzymes. However, only a few studies have examined the role of soil extracellular enzyme activities in the HFA of litter decomposition. Anthropogenic N input, including that of NHx and NOx, results in a proportional increase in terrestrial N addition (Hinkel et al., 2015). Many studies have demonstrated that N addition significantly influences the activity of microorganism and soil fauna during decomposition (Bardgett et al., 2013; Gan et al., 2013; Amend et al., 2015). The effects of N addition on soil communities will vary based on ecosystem, litter quality, and decomposition (Meunier et al., 2016). N addition could be beneficial to decomposition through the following: (1) N addition stimulates the activities of microorganisms; the input of extra N meets the N demands of microbes and accelerates the litter decomposition (Deng et al., 2007). (2) N addition increases the activity of enzymes; cellulase and amylase activities are limited by N content, so an extra N subsidy increases their activities and other glucosidase activities (Berg et al., 2000). (3) N addition increases plant biomass production and litter input, thereby decreasing litter C:N ratio (Henry et al., 2005). As often observed, the lower the litter C:N ratio is, the faster it decomposes (Taylor et al., 1989; Aerts, 1997; Ge et al., 2013). However, the added N may also polymerize with certain substances, such as polyphenol, produced in decomposition to form recalcitrant substances, thereby increasing litter lignin content and reducing litter decomposition rates (Aerts et al., 2006; Manning et al., 2008). Although the effects of N addition on decomposition have been studied, only a few studies have examined the effects of N addition on the HFA of litter decomposition (Vivanco and

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Austin, 2011; Allison et al., 2013). Thus, studying what affects the occurrence, direction, and magnitude of HFA from a diverse number of ecosystems during litter decomposition is necessary. Currently, conclusions on the direction and magnitude of HFA are not consistent. Previous studies on HFA mainly focus on single species litter in mono-specific forests or monoculture plantations (Mayor and Henkel, 2006; Barlow et al., 2007; Bachega et al., 2016). Furthermore, whether activities of specialized decomposers associated with the dominant species litter would be affected or not by environmental factors, such as N addition remains unclear. To better understand the HFA effect, we performed a field reciprocal litter transplant using mesh litterbags under different N-added treatments in a Chinese subtropical coniferous forest (dominated by Pinus massoniana) and broad-leaved forest (dominated by Quercus variabilis). The main objective of this study was to explore the effects of litter quality, habitat (home vs away) and soil communities (microorganisms and mesofauna) on decomposition rate depending on subtropical broad-leaved and coniferous forests. In addition to the HFA measurement for Q. variabilis and P. massoniana litter, we investigated the effect of N addition on the HFA with a particular focus on whether the specialized decomposer community that leads to a HFA of litter decomposition could be influenced. Because N addition improves the nutrient status of the ecosystem that with poor soil nutrition and promotes specialized decomposers activities, we expect that N addition would enhance the HFA effect in the low-nutrient coniferous forest. 2. Methods and materials 2.1. Study site This study was conducted in a 24 km2 subtropical forests on Zijin Mountain (447.1 m asl, 32°500 N, 118°480 E), Nanjing, Jiangsu, China from February to December 2014. The area has a subtropical monsoon climate. Mean annual rainfall is 1106.5 mm falling mostly between June and July, and mean annual air temperature is 15.4 °C (minimum of 1.9 °C in January and maximum of 28.2 °C in July). The soils are a slightly acidic humic cambisol with a pH of 5.0 ± 0.02 (Lv et al., 2014). The bedrock materials are sandstone and shale, and a large amount of nutrient and organic matter is deposited in the humus layer. The dominant forest types are deciduous broad-leaved forest and evergreen coniferous forest. In the broad-leaved forest, the dominant species is Q. variabilis with a relative basal area of approximately 85%, and that of P. massoniana is approximately 75% in coniferous forest (unpublished data of the field investigation 2011). 2.2. Experiment design In each forest, 30 plots of 1 m  1 m were established with 5 m between two adjacent plots. In half of the plots, an aqueous solution of NH4NO3 was added in equal doses each month at a rate of 47 kg N per hectare per year. Litter decomposition was measured in each forest using the litterbag method (Verhoef and Brussaard, 1990). In each plot, nylon litterbags (15 cm  20 cm) with two mesh sizes were used. A fine mesh of 0.2 mm was assumed to selectively impede the activity of the whole soil biota except microorganisms. A coarse mesh of 2 mm was assumed to allow microorganisms plus mesofauna activity. In October and November 2013, freshly senesced leaves of P. massoniana, Q. variabilis, were collected and air-dried for one month to a constant weight. Then, 6 g (±0.1) dry weight of each species litter was added to the litterbags. In February 2014, litterbags were placed in plots. Before the placement of litterbags, the litter layer and humus layers were

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removed so that the litterbags could be in direct contact with the mineral soil. Wooden screws were used to affix the litterbags in position. A total of 240 litterbags (2 mesh sizes  2 litter species  2 treatments  3 replicates  2 forest types  5 collection times) were used. The litterbags were collected for analysis at months 2, 4, 6, 8, and 10. In total, 48 litterbags were collected into polyethylene bags at each collection time, and soil samples just below the litterbags were collected using a metal 5 cm diameter soil corer. In the laboratory, the litter samples were oven-dried at 60 °C for one week to a constant weight to determine the mass remaining. Litter leaves were wiped thoroughly before weighing to prevent soil contamination. Soil samples were sieved through 2 mm mesh prior to soil pH, soil chemical properties, soil microbial biomass, and enzyme activity analyses.

where Xt is the litter mass remaining (g) at time t (month), X0 is the initial mass of the litter, and k is the litter decomposition rate constant (month–1). The HFA index (HFAI) for litter reciprocal transplant decomposition in the broad-leaved and coniferous forests was calculated in two steps based on relationships according to Ayres et al. (2009). First, the relative mass loss was calculated as

2.3. Mesofauna extraction

HFA index ¼

Mesofauna, mainly mites and springtails, were extracted using Tullgren funnels (Crossley and Blair, 1991) from the litter samples of each bag for one week post-collection. Organisms were stored in 70% alcohol, counted using a stereoscopic microscope, and identified to genus Collembola and order Acari (Faber, 1991; Lavelle, 1996; Chomel et al., 2015). Other invertebrates were assigned to functional groups according to Moore et al. (2005).

HFAI stands for the additional decomposition at home versus away habitat and is a net value for both species (A and B) in the reciprocal transplant. We used a two-sided, one-sample t-test to determine if the HFAI was significant. The relationship between the response variables (e.g., decomposition rate) and litter (broad-leaved vs. needle-leaved) and habitat (broad-leaved forest vs. coniferous forest) with interaction was determined using ANOVA. The mean HFA (% increase in k value at home compared with the away habitat) for each litter type was calculated based on the work of Austin et al. (2014):

2.4. Chemical analysis, soil pH, soil microbial biomass, and enzyme activity determination Total C and N concentrations of dried litter and soil samples were determined using an elemental analyzer (Elemental Vario Micro, Germany). The lignin concentration of the litter samples was measured using a gravimeter applying hot sulfuric acid digestion (Osono and Takeda, 2002). Soil microbial biomass was measured using the substrate-induced respiration (SIR) method (Osono and Takeda, 2002). All soil samples were controlled at 60% dry weight to remove any potential water limitation. Subsamples of fresh soil (1 g) were placed in a glass vial (100 mL). Next, 1 ml of an aqueous glucose solution (10 mg glucose g–1 soil) was added to each vial. Then samples were sealed and incubated at 25 °C for 1 h. CO2 produced by soil microbes was determined using an infrared gas analyzer (Bailey et al., 2002). Soil pH was measured using a glass electrode 1:2.5 (soil to water) ratio after shaking for approximately 30 min (Dick et al., 2000). Extracellular enzymes responsible for C cycling (cellobiohydrolase, CBH1; b-1,4-glucosidase, BG; and b-1,4-xylosidase, BX), N cycling (nitrate reductase, NR and URE), and P cycling (acid phosphatase, ACP; alkaline phosphatase, ALP) were determined using spectrophotometer. Soil samples were stored in a refrigerator at 4 °C before enzyme activity was measured. For CBH1, BG, BX analyses a modified method of Vepsalainen et al. (2001) was used. A standard curve was developed using PNPX. For NR and URE the method of Daniel and Curran (1981) and Nannipieri et al. (1980) were used and a standard curve was developed using KNO3 and urea. For ACP, ALP a modified method of Kandeler et al. (1999) was used and a standard curve was developed using disodium phenyl phosphate. The detailed methods are described in supplemental material. 2.5. Data analyses Mass remaining was expressed as the percentage of the total initial dry mass. The litter decomposition rate was determined using the following exponential equation (Olson, 1963):

Xt ¼ X0 ekt ;

ARMLa ¼

Aa  100 Aa þ Ba

ð1Þ

where ARMLa represents the relative mass loss of the litter from species A at habitat a, and Aa and Ba represent the percentage (of initial) litter mass loss of plant species A and B decomposing at habitat a. The HFAI was then calculated as

ARMLa þ BRMLb  100  100 ARMLb þ BRMLa

Mean HFA ¼ ðkhome  kaway Þ=kaway  100

ð2Þ

ð3Þ

where khome and kaway are the decomposition constants of a given species at home and in away habitats, respectively. All statistical tests were performed using SPSS (Version 19.0). Data were checked for deviations from normality and homogeneity of variance before analysis. ANOVAs and Tukey’s honestly significant differences tests were applied to assess significant differences between the various treatments. Multi-way ANOVAs were used to examine effects of treatment, mesh size, habitat and litter species on litter decomposition rate. Multi-way ANOVAs were also used to compare the magnitude of mean HFA among mesh size, litter species and treatment. Repeated-measures ANOVA were used to determine the effects of sampling time (repeated factor), treatment, habitat on soil analysis of chemical properties, biota and enzyme activities. One-way ANOVA was used to compare the magnitude of initial chemical properties of litter. 3. Results 3.1. Decomposition rates and HFAI The mass remaining of Q. variabilis and P. massoniana litters decreased rapidly during the first eight months and slowly declined in the latter two months in all treatments (Fig. 1). Litter mass remaining was significantly different for both litter types (p = 0.005, F = 5.631) and mesh size (p < 0.001, F = 14.289). Overall, the decomposition of the two species litter was greater in the coarse-mesh than in the fine-mesh litterbags, especially after 8 months (Tukey test, p < 0.001) (Fig. 2, Table 2). In the control plots, during 10 months of decomposition in the coarse-mesh litterbags, the mass remaining of Q. variabilis was 39.1% in the home habitat and 43.8% in the away habitat. The mass remaining of P. massoniana was 16.1% in the home habitat and 17.5% in the away habitat. The differences of the two litters between the home and away habitats were not significant (p > 0.05).

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Fig. 1. Changes in mass remaining of Q. variabilis and P. massoniana litter during 10 months of litter decomposition in fine mesh and coarse mesh litterbags and control and N treatment. Error bars indicate standard error (n = 3). BF: Broad-leaved forest; CF: Coniferous forest.

N addition had a significant effect on litter decomposition among different species in coarse-mesh litterbags (Figs. 1 and 2, Table 2). After 10 months of incubation in N-added plots and coarse-mesh litterbags, the mass remaining of P. massoniana was lower in the home than in the away habitat: 6.0% in coniferous forest vs. 29.9% in broad-leaved forest (Tukey test, p < 0.05). In addition, the decomposition rate (mean k value) of P. massoniana litter was 0.36 in coniferous forest (home) vs. 0.18 in broadleaved forest (away) (p < 0.05, Fig. 2). In the fine-mesh bags, the mass remaining of both species was not significantly different between the home and away habitats in the control plots or in N-added plots. N addition decreased the decomposition rates of P. massoniana litter in the fine-mesh litterbags in coniferous forests, but the differences were not significant (Fig. 2). Decomposition rates of Q. variabilis showed no significant difference between the two habitats in either fine-mesh or coarse-mesh litterbags (Fig. 2). The HFAI of litter decomposition for the two dominant species in normal condition (control plots) was negative in both the coarse- and fine-mesh litterbags (Fig. 3). However, a significant HFA effect was observed in coarse-mesh decomposition under N addition (HFAI was 26.892, p = 0.044), but not in control plots. The mean HFA of Q. variabilis and P. massoniana litter in the fine-mesh litterbags was negative (Tukey test, p < 0.001) when decomposed in the two forests in both the N-added and control

treatment. Conversely, P. massoniana litter in the coarse-mesh litterbags under N-added in the coniferous forest showed a positive HFA, and the value of the mean HFA was the highest of the experiment (20.7, p < 0.001) (Fig. 4). 3.2. Changes in litter quality during decomposition After 10 months of the field experiment, the C:N ratios of Q. variabilis and P. massoniana litter, generally declined both in the control plots and N addition plots. In the control plots, the C:N ratio of Q. variabilis litter in fine-mesh litterbags decreased by 48.9% in the home forest and 51.6% in the away forest. The C:N ratios in the coarse-mesh litterbags decreased by 54.5% in the home forest and 53.5% in the away forest. The C:N ratios of P. massoniana in the fine-mesh litterbags decreased by 63.0% in the home forest and 50.9% in the away forest. Conversely, the C:N ratios in the coarse-mesh litterbags decreased by 48.2% in the home forest and 41.6% in the away forest. In the N addition plots, the C:N ratios of Q. variabilis litter in the fine-mesh litterbags decreased by 54.5% in the home forest and 50.8% in the away forest. In the coarse-mesh litterbags C:N ratios decreased by 57.4% in the home forest and 55.3% in the away forest. The C:N ratios of P. massoniana in the fine-mesh litterbags decreased by 56.8% in the home forest and 61.4% in the away forest. In the coarse-mesh litterbags C:N ratios decreased by 47.7% in the home forest and 51.3% in the away

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Fig. 2. Litter decomposition constants (k) of Q. variabilis and P. massoniana litter in fine and coarse mesh litterbags after 10 months of decomposition at ‘home’ or ‘away’ without (control) and with N addition. Error bars indicate standard error (n = 3). Bars with the different letter are significantly different at p < 0.05.

Table 1 Initial chemical properties of P. massoniana and Q. acutissima litter (from one-way ANOVA) and the soil characteristics after 10 months of decomposition in control and N-addition plots (from repeated-measures ANOVA).

Initial litter properties Total C (%) Total N (%) Lignin (%) C/N Lignin/N

Q. acutissima

P. massoniana

46.30a 0.75a 30.52b 61.73b 40.69b

31.97b 0.37b 41.20a 86.41a 111.35a

Soil characteristics (10 months)

BF-N addition

BF-Control

CF-N addition

CF-Control

pH Total C (%) Total N (%) C/N Substrate-induced respiration (SIR) (lL CO2 h1 g1 Soil) Soil fauna Density (individuals/m2) Shannon-Wiener index (H) Menhinick index

4.42ab 5.71a 0.36a 15.72b 31.09b 212b 1.56a 0.38a

4.38b 4.91b 0.32a 15.49b 29.89b 195b 1.15ab 0.31ab

4.48ab 4.61b 0.27b 16.69a 37.36ab 295a 0.92b 0.27b

4.62a 4.93a 0.3ab 16.44a 42.49a 217b 1.32ab 0.34ab

Data of soil analysis represent mean values of 10 months sampling of litter decomposition in field experiment. Data with different superscript letters (a and b) in row indicates significant difference (p<0.05, n=3). BF: Broad-leaved forest and CF: Coniferous forest.

H. Lin et al. / Forest Ecology and Management 398 (2017) 216–225 Table 2 Effects of habitat, mesh size, litter species, N addition and their interactions on the decomposition constant k after 10 months of decomposition.

Habitat Mesh Litter species N addition Mesh  N addition Mesh  litter species N addition  litter species Habitat  litter species Habitat  Mesh  N addition Habitat  N addition  litter species N addition  litter species  mesh

df

F

P

1 1 1 1 1 1 1 1 1 1 1

1.851 55.955 24.297 5.936 0.45 30.322 0.237 0.452 2.563 1.13 2.673

0.183 <0.001 <0.001 0.016 0.503 <0.001 0.913 0.502 0.04 0.349 0.042

Significant differences are indicated by bold-face P values.

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(r2 = 0.90, p < 0.001), but negatively correlated with the litter N content (r2 = 0.72, p = 0.004). Compared with the control plots, N addition resulted in a significant negative relationship between the litter decomposition rates of the fine-mesh litterbags and C:N ratios in the coniferous forest (r2 = 0.56, p = 0.02) (Table S1). 3.3. Soil enzyme activities in N addition and control treatments Soil enzyme activity varied with collection time, habitat, and N addition (Tables 3 and 4) during 10 months of litter decomposition. In control treatments, most soil enzyme activity responsible for C and N cycling were higher in the broad-leaved forest than in the coniferous forest (Table 3). In the broad-leaved forest, N addition increased enzyme activity except for URE. In the coniferous forest, N addition significantly inhibited most enzyme activity except NR. Furthermore, N addition reduced the dissimilarity of soil microbial biomass in both habitats. N addition significantly increased the ratio of ALP and ACP activity from 1.605 to 1.794 in the broad-leaved forest, and decreased the ratio of ALP and ACP activity from 1.765 down to 1.607 in the coniferous forest (Table 3). Pearson’s correlation analysis showed a negative relationship between most soil microbial enzyme activity and litter decomposition rates of fine-mesh litterbags in N-added plots. However, there were some notable exceptions: BX and BG activity were positively correlated with the decomposition of Q. variabilis in fine-mesh litterbags in broad-leaved forest (p = 0.042, 0.008), and CBH1 was positively correlated with Q. variabilis decomposition in fine-mesh litterbags in the coniferous forest (p = 0.001). No correlations were observed between enzyme activity and litter decomposition rates of the coarse-mesh litterbags in both the broad-leaved and coniferous forests. 3.4. Changes in soil fauna and microbes

Fig. 3. Parameter estimates (Mean ± SE, n = 3) calculated for HFA index. Bars with different letters are significantly different at p < 0.05. Error bars indicate standard error. N: N addition and C: Control.

Fig. 4. The mean home-field advantage (HFA) for Q. variabilis and P. massoniana litter mass loss in fine and coarse mesh litterbags in control and N addition treatments. Error bars indicate standard error (n = 3). Bars with the different letter are significantly different at p < 0.05.

forest. Litter decomposition rates of the coarse-mesh litterbags had no relationship with the litter C:N and N contents in the N-added plots of the broad-leaved forest, whereas the litter decomposition rates of the coarse-mesh litterbags in the N-added plots of coniferous forest were positively correlated with the litter C:N ratios

SIR was used to estimate soil microbial biomass (Bailey et al., 2002; Lipson, 2007). Repeated-measures ANOVA showed that SIR was influenced by time, habitat, and N addition (Tables 1 and 4). In the control plots, the SIR of the coniferous forest increased quickly initially to a maximum in the fourth month of decomposition and then slightly decreased toward the end of the experimental period (Fig. S1). SIR was significantly greater in the coniferous forest than the broad-leaved forest in control plots (Tukey test, p < 0.05) (Table 1). In N-addition plots, microbial respiration changes were positively related with decomposition rates in the fine mesh litterbags in the two forests. At the first four months of decomposition, N addition reduced the SIR in both the broadleaved and coniferous forests though the difference was only significant in the coniferous forest (Table 1). During the last six months of decomposition, N addition increased SIR in the broadleaved forest, but it had no effect in the coniferous forest (Fig. S1). Pearson’s correlation analysis showed that the decomposition rates of the two litters in the fine-mesh litterbags were positively correlated with SIR in the control plots of the two forests. In the coniferous forest, P. massoniana decomposition rates in the coarse-mesh litterbags were significantly negatively correlated with SIR in N addition plots (r2 = 0.36, p = 0.018) (Table S2). Mesofauna was dominated by Acari (proportion ranging between 28% and 61% of total abundance) and Collembola (proportion ranging between 2% and 45%) for all the plots (Fig. 5). In both control plots and N addition plots, Collembola abundance was significantly lower at month 8 during decomposition (Tukey test, p < 0.05); the abundance of detritivorous mites was significantly higher in coniferous forest litter than broad-leaved forest litter for all months of decomposition (Tukey test, p < 0.05), whereas the abundance of other predators and detritivores were lower only

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Table 3 Effect of N addition on extracellular enzyme activity after 10 months of incubation in the two forest types. Enzyme

NR(IU) ACP(IU) ALP(IU) URE(IU) BX(IU) BG(IU) CBH1(IU)

Broad-leaved forest

Coniferous forest

N addition

Control

N addition

Control

5.05 ± 0.58a 13.18 ± 1.68a 23.64 ± 7.78a 3.3 ± 0.60b 2.1 ± 0.09a 2.15 ± 0.03a 2 ± 0.10a

4.7 ± 0.45b 12.9 ± 1.60b 20.71 ± 4.49b 3.71 ± 0.61a 2.08 ± 0.04ab 2.06 ± 0.09b 1.89 ± 0.19b

4.26 ± 0.51b 12.7 ± 1.78b 20.42 ± 1.61b 3.42 ± 0.47b 2.04 ± 0.05b 2.1 ± 0.03a 1.97 ± 0.15b

3.7 ± 0.3c 13.11 ± 1.27a 23.15 ± 3.73a 3.51 ± 0.45ab 2.04 ± 0.08b 2.07 ± 0.08ab 2.04 ± 0.05a

Data represent mean values of 10 months sampling during litter decomposition and standard error (n = 3). Different superscript letters (a, b and c) in a row show, for two given forests, significant differences among treatments (p < 0.05). NR: nitrate reductase; ACP: acid phosphatase; ALP: alkaline phosphatase; URE: urease; BX: b-1,4xylosidase; BG: b-1,4-glucosidase; and CBH1: cellobiohydrolase.

Table 4 Effects (indicated by P values from repeated-measures ANOVA) of sampling date, habitat, N addition and their interactions on SIR and enzyme activity in broad-leaved and coniferous forest. Variation

SIR

PH

NR

ACP

ALP

URE

BX

BG

CBH1

Between subjects Intercept Treatment Habitat Treatment  Habitat

<0.001 0.035 <0.001 0.049

<0.001 0.475 0.05 0.226

<0.001 0.161 0.015 0.73

<0.001 0.926 0.858 0.038

<0.001 0.868 0.965 0.025

<0.001 0.117 0.78 0.289

<0.001 0.811 0.167 0.812

<0.001 0.05 0.587 0.375

<0.001 0.684 0.23 0.084

Within subjects Date Date  Treatment Date  Habitat Date  Treatment  Habitat

<0.001 0.076 0.263 0.085

<0.001 0.001 0.002 0.307

0.002 0.642 0.008 0.013

<0.001 0.558 0.068 0.239

<0.001 0.031 0.884 0.021

<0.001 0.005 <0.001 0.295

<0.001 0.051 0.984 0.001

<0.001 0.001 0.682 0.638

<0.001 0.717 0.008 <0.001

Significant P-values are presented in boldface. SIR: soil substrate-induced respiration; NR: nitrate reductase; ACP: acid phosphatase; ALP: alkaline phosphatase; URE: urease; BX: b-1,4-xylosidase; BG: b-1,4-glucosidase; and CBH1: cellobiohydrolase.

Fig. 5. Abundance dynamics of mesofauna functional groups present in litter from broad-leaved and coniferous forests with N addition and control treatments during 10 months of litter decomposition. BF: Broad-leaved forest; CF: Coniferous forest; N: N addition; and C: Control.

at month 6 (Tukey test, p < 0.05). After 10 months of decomposition, Collembola and Acari biomass had increased significantly in all plots (Tukey test, p < 0.01). In the coniferous forest, the abundance of mesofauna was significantly greater under N addition. 4. Discussion 4.1. Litter decomposition rates and HFA in broad-leaved and coniferous forests Our findings show that during decomposition, there are no significant differences in litter decomposition rates between home

and away forests for both P. massoniana and Q. variabilis. This suggests the absence of HFA effect in both coniferous and broadleaved forests. Several studies reported the absence of HFA effect because the habitat was not a significant factor affecting litter decomposition despite large differences in litter quality which is consistent with our findings (Giebelmann et al., 2011; St. John et al., 2011; Yu et al., 2015). However, by conducting a metaanalysis of published data in global forest ecosystems, Wang et al. (2012) suggested that the HFA in litter decomposition was widespread in forest ecosystems, showing that the litter mass loss was on average 4.2% faster in the home environment. As we know, low-quality, recalcitrant (high C:N, lignin:N) litter is difficult to

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decompose, and soil microbes adapt recalcitrant litter (Strickland et al., 2009b), such as phenolic compounds in coniferous forests. In broad-leaved forests, few soil microbes adapt to decompose the recalcitrant litter (Milcu and Manning, 2011; Austin et al., 2014; Chomel et al., 2015; Veen et al., 2015b). By contrast, the high-quality (low C:N, lignin:N) litter is relatively easy to decompose for most decomposers that can break down these compounds regardless of the habitat. However, we found that both P. massoniana and Q. variabilis litters did not decompose significantly faster in their home habitat than in the away habitat, which showed that although litter quality significantly affects litter decomposition, it may not be the main factor that drives the HFA effect. Soil community and abiotic factors should also be considered comprehensively. Firstly, although soil microbial biomass between the broad-leaved and coniferous forests changed after litter transplant, the abundance of soil fauna specialized for P. massoniana and Q. variabilis litter decomposition was not significantly different between the home and away forests. In addition, the suitable temperature and moisture conditions of our experiment maintained the diversity and abundance of soil fauna to decompose different litter quickly. Therefore, the decomposition of the two litter types was not different between the home and away habitats, resulting in a lack of HFA effect. Secondly, the microbial decomposers, especially fungus, likely have broad functions and flexible adaptability decomposing the two litters regardless of habitats (St. John et al., 2011; Chomel et al., 2015). Thus, changes in microbial community composition and biomass did not lead to changes in decomposition rates. The HFA effect of litter decomposition is also affected by the activity of soil decomposers during decomposition in the home and away habitats (Aerts and de Caluwe, 1997; Wang et al., 2012). In the early stages of decomposition, because fresh litters contain some water-soluble or easily decomposed substrate, microorganisms, mainly fungus, are most active. In the late stages of decomposition, the tougher to decompose compounds such as cellulose accounts for the major parts of the litter remaining. Leaching and microorganisms begin to weaken, and soil fauna with powerful feeding activity become largely involved (Xu et al., 2005a). Composition adjustment of soil community for litter decomposition of different stages could weaken the specialization of the decomposers and substrate. In the present study, the abundance of soil fauna increased after 6 months. This is possibly because of soil fauna grazing activities decreasing microbial biomass, especially for coniferous species (Chomel et al., 2015). Thus the specificity of microorganisms to liable litter could be weakened, which could also partly explain the lack of HFA effect. 4.2. Effects of N addition on HFA effect The greater decomposition rate of P. massoniana at home in the coniferous forest in N addition plots suggests that N addition enhanced the magnitude of HFA effect for P. massoniana. There are also some studies demonstrating that environmental factors increasing litter decomposition in coniferous forests. For example, Moore et al. (1999) found that elevated atmospheric CO2 increased the decomposition rates of 18 different litter types in five needleleaved forests in Canada. Zhou et al. (2014) indicated that higher temperatures could also accelerate litter decomposition by enhancing litter microbial metabolic activity in a Korean pine forests. The activity of microbial decomposers is significantly related litter decomposition rates. However, we found that the habitat specificity of soil fauna regardless of litter type is essential in generating the HFA effect of litter decomposition, which is consist with results from Nielsen’ study (2010a). In the present study, soil fauna abundance increased after 6 months of N addition in the coniferous forest, which accelerated the decomposition rates of the recalcitrant

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P. massoniana litter in coarse-mesh litterbags in the home habitat. The N contents of both coniferous soil and P. massoniana litter are low. N may be the limiting factor in the ecosystem function. N addition may meet the N demands for the ecosystem (Lü et al., 2013; Lv et al., 2013). As a result, in the coniferous forest, N addition improved the nutrient status and promoted soil fauna activities and litter decomposition rate. The greater soil fauna abundance in the coniferous forest, could be explained by a strong habitat specificity due to their weak diffusion ability (Nielsen et al., 2010b). Milcu and Manning (2011) supported the idea that soil biota were responsible for the HFA of decomposition. They found the HFA effect was mainly due to accelerated decomposition caused by the specialization of mesofauna. For the above reasons, N addition significantly increased the soil fauna numbers and activity in the coniferous forest, thereby strengthening the specificity effects of soil fauna for decomposing a particular litter type or particular compound, and causing the positive HFA effect on P. massoniana litter decomposition but negative effect on Q. variabilis. Xu et al. (2005b) found the diversity of soil fauna was increased significantly in a P. massoniana-dominated forest with 16 months of N addition, thereby promoting P. massoniana litter decomposition, which is in agreement with our findings. In addition, litter decomposition rate and HFA effect are associated with soil fertility (Hilli et al., 2010; Ge et al., 2013). The higher the C:N ratio of the soil could result in greater litter decomposition rate and nutrient release (Bonanomi et al., 2016). For example, Aerts and de Caluwe (1997) showed that the litter of the Carex species decomposed in poor soil released nutrients more quickly than that in fertile soil. In the present study, the soil C:N ratios were higher in the coniferous forest, which may explain why the P. massoniana and Q. variabilis litters decomposed faster in coniferous forests. It has been suggested that microbial decomposers, such as saprophytes specialize in the decomposition of a certain substrate (Gießelmann et al., 2011). Furthermore, Rühling and Tyler (1991) and Newell et al. (1996) found that N addition increased the biomass and abundance of saprophytes. The specificity and diversity of microbial decomposers between the home and away habitats should have led to an HFA effect (Prescott and Grayston, 2013), but our findings contradicted this expected result. We showed N addition reduced the SIR and most enzyme activity in the coniferous forest, which fit with previous results reported for SIR responses to N additions. For example, Janssens et al. (2010) using a meta-analysis of measurements in N-addition experiments, showed that the negative effect of N on soil respiration is widespread. Carriero and Frey (Carreiro et al., 2000; Frey et al., 2004) detected added N inhibited the activity of the lignin-degrading enzyme, phenol oxidase produced by white-rot fungi. That is to say, N addition has potentially negative effects on priming effects in soil (Blagodatskaya and Kuzyakov, 2008). A possible reason, especially in the coniferous forest, may be the ectomycorrhizal fungal community being quite sensitive to N inputs (Bowden et al., 2004). It is also possible that decreases in soil pH reduce microbial activity (Aerts and Caluwe, 1999). In both forests, mineral soil pH measured lower in the fertilized plots than in control plots. In the coniferous forest, control pH was 4.62, and the N plot was 4.48. Furthermore, due to the short life cycle of soil bacteria and the rapid reproductive ability of fungus, microbial decomposers could shift their community structure within a short period to adapt to the decomposition of different substrates and N addition. Previous studies also demonstrate the flexibility and dynamics of microbial communities and the functional redundancy of the decomposer community (Giebelmann et al., 2011; Yu et al., 2015). The functional adjustment of the microbial community to N addition could weaken the HFA effect in litter decomposition of the fine-mesh litterbags in the broad-leaved and coniferous forests. Soil microbial

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activities are susceptible to the activities of the soil fauna community. We found no positive relationships between the litter decomposition rates and soil enzyme activity in the coarse-mesh litterbags. This result could be the effect of grazing activity of fungivores (such as mites and collembola) reducing hyphal growth and decreasing microbial biomass, thereby reducing soil enzymes, such as phosphatase, URE, and cellulase. Furthermore, soil microorganisms, such as fungus, bacteria, and actinomycetes, are important in litter enzymolysis (Dilly and Munch, 1996; Bandick and Dick, 1999; Groffman et al., 2001). Soil enzymes express specificity to break down litter substrates (Chomel et al., 2015). Our findings indicate that under N-addition, the soil enzymes responsible for C cycling (BX and BG) promote Q. variabilis litter decomposition in the broad-leaved forest. Also, CBH1 involved in C cycling promoted Q. variabilis litter decomposition in the coniferous forest. The specificity for substrate of soil enzymes, although not significantly related to habitat, could explain the lack of significant difference of Q. variabilis litter decomposition between the home and away habitats. However, no soil enzymes affect P. massoniana litter decomposition, such as that of Q. variabilis litter, between the home and away habitats. Moreover, the ratio of ALP and ACP activity is a sensitive indicator of the function of microbial decomposition (Burns, 1982; Dick et al., 2000). Our findings showed that N addition had no significant effect on most soil-enzyme activity and the ratio of ALP and ACP activity in the coniferous forest, which suggesting that microorganisms may not be the main drivers of the HFA effect. In summary, no HFA effect was observed for the decomposition of the dominant litter species in the subtropical forests. However, we found a positive HFA effect of P. massoniana litter decomposition under N addition compared to the control plots, in which the soil mesofauna are likely the most important factor due to their greater diversity and abundance in the home habitat compared to the away habitat. Acknowledgments This study was financially supported by the National Key Research and Development Program of the Ministry of Science and Technology of China (No. 2016YFD0600204) and the State Key Program of National Natural Science Foundation of China (No. 3153007). We would also like to thank Alison Beamish at the University of British Columbia for her assistance with English language and grammatical editing of the manuscript. Appendix A. Supplementary material Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.foreco.2017.05. 015. References Aerts, R., 1997. Climate, leaf litter chemistry and leaf litter decomposition in terrestrial ecosystems: a triangular relationship. Oikos 79, 439–449. Aerts, R., Caluwe, H.D., 1999. Nitrogen deposition effects on carbon dioxide and methane emissions from temperate peatland soils. Oikos 84, 44–54. Aerts, R., de Caluwe, H., 1997. Nutritional and plant-mediated controls on leaf litter decomposition of Carex species. Ecology 78, 244–260. Aerts, R., van Logtestijn, R.S., Karlsson, P.S., 2006. Nitrogen supply differentially affects litter decomposition rates and nitrogen dynamics of sub-arctic bog species. Oecologia 146, 652–658. Allison, S.D., Lu, Y., Weihe, C., Goulden, M.L., Martiny, A.C., Treseder, K.K., Martiny, J. B., 2013. Microbial abundance and composition influence litter decomposition response to environmental change. Ecology 94, 714–725. Amend, A.S., Matulich, K.L., Martiny, J.B., 2015. Nitrogen addition, not initial phylogenetic diversity, increases litter decomposition by fungal communities. Front. Microbiol. 6.

Austin, A.T., Vivanco, L., Gonzalez-Arzac, A., Perez, L.I., 2014. There’s no place like home? An exploration of the mechanisms behind plant litter-decomposer affinity in terrestrial ecosystems. New Phytol. 204 (2), 307–314. Ayres, E., Steltzer, H., Simmons, B.L., Simpson, R.T., Steinweg, J.M., Wallenstein, M.D., Mellor, N., Parton, W.J., Moore, J.C., Wall, D.H., 2009. Home-field advantage accelerates leaf litter decomposition in forests. Soil Biol. Biochem. 41, 606–610. Bachega, L.R., Bouillet, J.-P., de Cássia Piccolo, M., Saint-André, L., Bouvet, J.-M., Nouvellon, Y., de Moraes Gonçalves, J.L., Robin, A., Laclau, J.-P., 2016. Decomposition of Eucalyptus grandis and Acacia mangium leaves and fine roots in tropical conditions did not meet the Home Field Advantage hypothesis. For. Ecol. Manage. 359, 33–43. Bailey, V.L., Peacock, A.D., Smith, J.L., Bolton, H., 2002. Relationships between soil microbial biomass determined by chloroform fumigation–extraction, substrateinduced respiration, and phospholipids fatty acid analysis. Soil Biol. Biochem. 34, 1385–1389. Bandick, A.K., Dick, R.P., 1999. Field management effects on soil enzyme activities. Soil Biol. Biochem. 31, 1471–1479. Bardgett, R.D., Manning, P., Morriën, E., Vries, F.T., 2013. Hierarchical responses of plant–soil interactions to climate change: consequences for the global carbon cycle. J. Ecol. 101, 334–343. Barlow, J., Gardner, T.A., Ferreira, L.V., Peres, C.A., 2007. Litter fall and decomposition in primary, secondary and plantation forests in the Brazilian Amazon. For. Ecol. Manage. 247, 91–97. Berg, B., Johansson, M., Meentemeyer, V., 2000. Litter decomposition in a transect of Norway spruce forests: substrate quality and climate control. Can. J. For. Res. 30, 1136–1147. Blagodatskaya, E., Kuzyakov, Y., 2008. Mechanisms of real and apparent priming effects and their dependence on soil microbial biomass and community structure: critical review. Biol. Fertil. Soils 45, 115–131. Bonanomi, G., Cesarano, G., Gaglione, S.A., Ippolito, F., Sarker, T., Rao, M.A., 2016. Soil fertility promotes decomposition rate of nutrient poor, but not nutrient rich litter through nitrogen transfer. Plant Soil 412, 397–411. Bowden, R.D., Davidson, E., Savage, K., Arabia, C., Steudler, P., 2004. Chronic nitrogen additions reduce total soil respiration and microbial respiration in temperate forest soils at the Harvard Forest. For. Ecol. Manage. 196, 43–56. Burns, R.G., 1982. Enzyme activity in soil: location and possible role in microbial ecology. Soil Biol. Biochem. 14, 423–427. Carreiro, M.M., Sinsabaugh, R.L., Repert, D.A., Parkhurst, D.F., 2000. Microbial enzyme shifts explain litter decay responses to simulated nitrogen deposition. Ecology 81, 2359–2365. Cesarz, S., Fender, A.-C., Beyer, F., Valtanen, K., Pfeiffer, B., Gansert, D., Hertel, D., Polle, A., Daniel, R., Leuschner, C., Scheu, S., 2013. Roots from beech (Fagus sylvatica L.) and ash (Fraxinus excelsior L.) differentially affect soil microorganisms and carbon dynamics. Soil Biol. Biochem. 61, 23–32. Chomel, M., Guittonny-Larchevêque, M., DesRochers, A., Baldy, V., 2015. Home field advantage of litter decomposition in pure and mixed plantations under boreal climate. Ecosystems 18, 1014–1028. Cleveland, C.C., Townsend, A.R., Taylor, P., Alvarez-Clare, S., Bustamante, M.M., Chuyong, G., Dobrowski, S.Z., Grierson, P., Harms, K.E., Houlton, B.Z., Marklein, A., Parton, W., Porder, S., Reed, S.C., Sierra, C.A., Silver, W.L., Tanner, E.V., Wieder, W.R., 2011. Relationships among net primary productivity, nutrients and climate in tropical rain forest: a pan-tropical analysis. Ecol. Lett. 14, 939–947. Crossley, D.A., Blair, J.M., 1991. A high-efficiency, ‘‘low-technology” Tullgren-type extractor for soil microarthropods. Agr. Ecosyst. Environ. 34, 187–192. Daniel, R.M., Curran, M.P., 1981. A method for the determination of nitrate reductase. J. Biochem. Biophys. Meth. 4, 131. Deng, X., Zhang, Y., Han, S., Li, X., Li, K., Li, X., 2007. Effects of nitrogen addition on the early stage decomposition of Korean pine litter in the Changbaishan Mountains, northeastern China. J. Beijing Forest. Univ. 29, 16–22. Dick, W.A., Cheng, L., Wang, P., 2000. Soil acid and alkaline phosphatase activity as pH adjustment indicators. Soil Biol. Biochem. 32, 1915–1919. Dilly, O., Munch, J.P., 1996. Microbial biomass content, basal respiration and enzyme activities during the course of decomposition of leaf litter in a black alder (Alnus glutinosa (L.) Gaertn.) forest. Soil Biol. Biochem. 28, 1073–1081. Faber, J.H., 1991. Functional classification of soil fauna: a new approach. Oikos 62 (1), 110–117. Frey, S.D., Knorr, M., Parrent, J.L., Simpson, R.T., 2004. Chronic nitrogen enrichment affects the structure and function of the soil microbial community in temperate hardwood and pine forests. For. Ecol. Manage. 196, 159–171. Gan, H., Zak, D.R., Hunter, M.D., 2013. Chronic nitrogen deposition alters the structure and function of detrital food webs in a northern hardwood ecosystem. Ecol. Appl. 23, 1311–1321. Ge, X., Zeng, L., Xiao, W., Huang, Z., Geng, X., Tan, B., 2013. Effect of litter substrate quality and soil nutrients on forest litter decomposition: a review. Acta Ecol. Sin. 33, 102–108. Gholz, H.L., Wedin, D.A., Smitherman, S.M., Harmon, M.E., Parton, W.J., 2000. Longterm dynamics of pine and hardwood litter in contrasting environments: toward a global model of decomposition. Glob. Change Biol. 6, 751–765. Gießelmann, U.C., Martins, K.G., Brändle, M., Schädler, M., Marques, R., Brandl, R., 2011. Lack of home-field advantage in the decomposition of leaf litter in the Atlantic Rainforest of Brazil. Appl. Soil. Ecol. 49, 5–10. Groffman, P.M., McDowell, W.H., Myers, J.C., Merriam, J.L., 2001. Soil microbial biomass and activity in tropical riparian forests. Soil Biol. Biochem. 33, 1339–1348. Hättenschwiler, S., Tiunov, A.V., Scheu, S., 2005. Biodiversity and litter decomposition in terrestrial ecosystems. Annu. Rev. Ecol. Evol. Syst. 36, 191– 218.

H. Lin et al. / Forest Ecology and Management 398 (2017) 216–225 Henry, H.A.L., Cleland, E.E., Field, C.B., Vitousek, P.M., 2005. Interactive effects of elevated CO2, N deposition and climate change on plant litter quality in a California annual grassland. Oecologia 142, 465–473. Hilli, S., Stark, S., Derome, J., 2010. Litter decomposition rates in relation to litter stocks in boreal coniferous forests along climatic and soil fertility gradients. Appl. Soil. Ecol. 46, 200–208. Hinkel, J., Jaeger, C., Nicholls, R.J., Lowe, J., Renn, O., Peijun, S., 2015. Sea-level rise scenarios and coastal risk management. Nat. Climate Change 5, 188–190. Janssens, I.A., Dieleman, W.I., Luyssaert, S.E.A., 2010. Reduction of forest soil respiration in response to nitrogen deposition. Nat. Geosci. 3, 315–322. Kandeler, E., Tscherko, D., Spiegel, H., 1999. Long-term monitoring of microbial biomass, N mineralisation and enzyme activities of a Chernozem under different tillage management. Biol. Fertil. Soils 28 (4), 343–351. Lü, Y., Wang, C., Jia, Y., Du, J., Ma, X., Wang, W., Pu, G., Tian, X., 2013. Responses of soil microbial biomass and enzymatic activities to different forms of organic nitrogen deposition in the subtropical forests in East China. Ecol. Res. 28, 447– 457. Lavelle, P., 1996. Diversity of soil fauna and ecosystem function. Biology Int. 33. Lipson, D.A., 2007. Relationships between temperature responses and bacterial community structure along seasonal and altitudinal gradients. FEMS Microbiol. Ecol. 59, 418–427. Lv, Y., Wang, C., Jia, Y., Wang, W., Ma, X., Du, J., Pu, G., Tian, X., 2014. Effects of sulfuric, nitric, and mixed acid rain on litter decomposition, soil microbial biomass, and enzyme activities in subtropical forests of China. Appl. Soil. Ecol. 79, 1–9. Lv, Y., Wang, C., Wang, F., Zhao, G., Pu, G., Ma, X., Tian, X., 2013. Effects of nitrogen addition on litter decomposition, soil microbial biomass, and enzyme activities between leguminous and non-leguminous forests. Ecol. Res. 28, 793–800. Manning, P., Saunders, M., Bardgett, R.D., Bonkowski, M., Bradford, M.A., Ellis, R.J., Kandeler, E., Marhan, S., Tscherko, D., 2008. Direct and indirect effects of nitrogen deposition on litter decomposition. Soil Biol. Biochem. 40, 688–698. Mayor, J.R., Henkel, T.W., 2006. Do ectomycorrhizas alter leaf-litter decomposition in monodominant tropical forests of Guyana? New Phytol. 169, 579–588. Meunier, C.L., Gundale, M.J., Sánchez, I.S., Liess, A., 2016. Impact of nitrogen deposition on forest and lake food webs in nitrogen-limited environments. Glob. Change Biol. 22, 164–179. Milcu, A., Manning, P., 2011. All size classes of soil fauna and litter quality control the acceleration of litter decay in its home environment. Oikos 120, 1366–1370. Moore, J.C., McCann, K., de Ruiter, P.C., 2005. Modeling trophic pathways, nutrient cycling, and dynamic stability in soils. Pedobiologia 49, 499–510. Moore, T.R., T, J.A., Taylor, B., et al., 1999. Litter decomposition rates in Canadian forests. Glob. Change Biol. 5, 75–82. Nannipieri, P., Ceccanti, B., Cervelli, S., Matarese, E., 1980. Extraction of phosphatase, urease, proteases, organic carbon, and nitrogen from soil. Soil Sci. Soc. Am. J. 44, 1011–1016. Naseby, D.C., Pascual, J.A., Lynch, J.M., 2000. Effect of biocontrol strains of Trichoderma on plant growth, Pythium ultimum populations, soil microbial communities and soil enzyme activities. J. Appl. Microbiol. 88, 161–169. Newell, S.Y., Arsuffi, T.L., Palm, L.A., 1996. Misting and nitrogen fertilization of shoots of a saltmarsh grass: effects upon fungal decay of leaf blades. Oecologia 108, 495–502. Nielsen, U.N., Osler, G.H., Campbell, C.D., Burslem, D.F., van der Wal, R., 2010a. The influence of vegetation type, soil properties and precipitation on the composition of soil mite and microbial communities at the landscape scale. J. Biogeogr. 37, 1317–1328. Nielsen, U.N., Osler, G.H.R., Campbell, C.D., Burslem, D.F.R.P., Der Wal, R.V., 2010b. The influence of vegetation type, soil properties and precipitation on the composition of soil mite and microbial communities at the landscape scale. J. Biogeogr. 37, 1317–1328. Olson, J.S., 1963. Energy storage and the balance of producers and decomposers in ecological systems. Ecology 44, 322–331.

225

Osono, T., Takeda, H., 2002. Comparison of litter decomposing ability among diverse fungi in a cool temperate deciduous forest in Japan. Mycologia 94, 421–427. Pfeiffer, B., Fender, A.-C., Lasota, S., Hertel, D., Jungkunst, H.F., Daniel, R., 2013. Leaf litter is the main driver for changes in bacterial community structures in the rhizosphere of ash and beech. Appl. Soil. Ecol. 72, 150–160. Prescott, C.E., 2005. Do rates of litter decomposition tell us anything we really need to know? For. Ecol. Manage. 220, 66–74. Prescott, C.E., Grayston, S.J., 2013. Tree species influence on microbial communities in litter and soil: current knowledge and research needs. For. Ecol. Manage. 309, 19–27. Rühling, Å., Tyler, G., 1991. Effects of simulated nitrogen deposition to the forest floor on the macrofungal flora of a beech forest. Ambio, 261–263. St. John, M.G., Orwin, K.H., Dickie, I.A., 2011. No ‘home’ versus ‘away’ effects of decomposition found in a grassland–forest reciprocal litter transplant study. Soil Biol. Biochem. 43, 1482–1489. Strickland, M.S., Lauber, C.L., Fierer, N., Bradford, M.A., 2009a. Testing the functional significance of microbial community composition. Ecology 90, 441–451. Strickland, M.S., Osburn, E., Lauber, C., Fierer, N., Bradford, M.A., 2009b. Litter quality is in the eye of the beholder: initial decomposition rates as a function of inoculum characteristics. Funct. Ecol. 23, 627–636. Taylor, B.R., Parkinson, D., Parsons, W.J., 1989. Nitrogen and lignin content as predictors of litter decay rates: a microcosm test. Ecology 70, 97–104. Ushio, M., Balser, T.C., Kitayama, K., 2012. Effects of condensed tannins in conifer leaves on the composition and activity of the soil microbial community in a tropical montane forest. Plant Soil 365, 157–170. Veen, G.F., Sundqvist, M.K., Wardle, D.A., Briones, M.J., 2015a. Environmental factors and traits that drive plant litter decomposition do not determine home-field advantage effects. Funct. Ecol. 29, 981–991. Veen, G.F.C., Freschet, G.T., Ordonez, A., Wardle, D.A., 2015b. Litter quality and environmental controls of home-field advantage effects on litter decomposition. Oikos 124, 187–195. Vepsalainen, M., Kukkonen, S., Vestberg, M., Sirvio, H., Niemi, R.M., 2001. Application of soil enzyme activity test kit in a field experiment. Soil Biol. Biochem. 33, 1665–1672. Verhoef, H.A., Brussaard, L., 1990. Decomposition and nitrogen mineralization in natural and agroecosystems: the contribution of soil animals. Biogeochemistry 11, 175–211. Vivanco, L., Austin, A.T., 2011. Nitrogen addition stimulates forest litter decomposition and disrupts species interactions in Patagonia, Argentina. Glob. Change Biol. 17, 1963–1974. Wallenstein, M.D., Haddix, M.L., Ayres, E., Steltzer, H., Magrini-Bair, K.A., Paul, E.A., 2013. Litter chemistry changes more rapidly when decomposed at home but converges during decomposition–transformation. Soil Biol. Biochem. 57, 311– 319. Wang, Q., Zhong, M., He, T., 2012. Home-field advantage of litter decomposition and nitrogen release in forest ecosystems. Biol. Fertil. Soils 49, 427–434. Xu, G., Mo, J., Zhou, G., Xue, J., 2005a. Litter decomposition under N deposition in Dinghushan forests and its relationship with soil fauna. Ecol. Environ. 14, 901– 907. Xu, G., Zhou, X., Zhou, G., 2005b. Responses of aboveground soil fauna community to simulated N deposition addition in forest ecosystems. Acta Scientiarum Naturalium Universitatis Sunyatseni 44, 213–221. Yu, Z., Huang, Z., Wang, M., Liu, R., Zheng, L., Wan, X., Hu, Z., Davis, M.R., Lin, T.-C., 2015. Nitrogen addition enhances home-field advantage during litter decomposition in subtropical forest plantations. Soil Biol. Biochem. 90, 188– 196. Zhang, Y.-M., Guo-yi, Z., Ning, W., 2004. A review of studies on soil enzymology. J. Trop. Subtrop. Bot. 12, 83–90. Zhou, Y., Clark, M., Su, J., Xiao, C., 2014. Litter decomposition and soil microbial community composition in three Korean pine (Pinus koraiensis) forests along an altitudinal gradient. Plant Soil 386, 171–183.