Introduction
Plant properties directly affect the productivity and function of an
ecosystem in a natural environment (Chapin et al., 1986; Chapin, 2003;
Berendse and Aerts, 1987; Grime, 1998). Plants continuously lose N and P in
their entire life history and even during litter production and decomposition
(Laungani and Knops, 2009; Richardson et al., 2009). In cold environments,
litter tends to be recalcitrant (Aerts, 1997), but reproductive tissues
present chemical composition that differs from vegetative parts, resulting in
a markedly faster decomposition and nutrient release, with repercussions for
nutrient cycling and patchiness (Buxton and Marten, 1989; Lee et al., 2011).
Although inflorescences comprise only a small fraction of plant biomass and
production in Arctic and alpine vegetation, the inflorescence production can
be a significant proportion of the total production of species under certain
special circumstances (Martínez-Yrízar et al.,
1999; Fabbro and Körner, 2004;
Wookey et al., 2009). High contents of N and P exist in the reproductive
organs of plants, probably because of their essential roles in plant growth
and formation (e.g., high protein content). The rate of decay and
concentrations of nutrients in the litter determine the rate of nutrient
release, which creates a positive feedback to site fertility. Hence, the
chemical properties of litters from different plant organs and their
correlations with decomposition rate must be determined.
The growth and health of plants in their life history have been considerably
influenced by variations in the physical, chemical, and biological
properties of soil, particularly around the rhizosphere. Nonetheless, soil
properties can also be mediated by plants. N is a major constituent of
several important plant substances (Vitousek and Howarth, 1991). In cold
life zone ecosystems, plant biomass production is limited by N (Körner,
2003). The plant residue is one principal component of soil organic matter
(SOM), whose decomposition can supply available N to plants and
microorganisms. Similar to N, P is closely associated with numerous vital
plant processes. Nevertheless, in most circumstances, P is limited because
of its small concentration in soil; this element is released slowly from
insoluble P but is highly demanded by plants and microorganisms (Bieleski,
1973; Richardson et al., 2009). As decomposition is a prolonged process,
plants contain concentrated nutrients comparable with soil, which enhance
the microbial immobilization of N when they provide C to soil
microorganisms. The nature of litter determines its palatability to soil
organisms, thereby influencing their composition and activity levels. Litter
can also mediate the interactions between neighboring plants in infertile
communities (Nilsson et al., 1999; Xiong and Nilsson, 1999), which
significantly affect the biogeochemical cycle and feedback of plant–soil
interaction.
Fast decay of N-rich litters suggests that litter decay rates increase with
increasing N content. The initial rate of nutrient release is positively
correlated with the initial concentrations of N or P (MacLean and Wein, 1978;
Aber and Melillo, 1980; Berg and
Ekbohm, 1983; Yavitt and Fahey, 1986; Stohlgren, 1988). Long-term increases
in N availability have also been reported following the additions of C to
forests (Groffman, 1999). In agricultural systems, addition of fresh residues
can stimulate the decomposition and net release of N from indigenous SOM
(Haynes, 1986; Scott et al., 1996). Recently, a common-garden decomposition
experiment in a wide range of subarctic plant types demonstrated that
structural and chemical traits are better predictors of several high-turnover
organs than structural traits alone (Freschet et al., 2012). Decomposition
rates of plant litters slightly differ because of their species-specific
traits and various organs, whose chemical qualities vary in a wide range of
plant types and environments.
Location of the study sites.
Alpine ecosystems are thermally restricted and characterized by a low
material turnover rate (Körner, 2003). In a high-altitude region, plants
grow in a harsh habitat that restricted their effective utilization of
resources; in this regard, the total available resource is less compared with
that of plants in other regions (Fabbro and Körner, 2004; Hautier et al.,
2009). In long-term evolution, the allocation of accumulated carbohydrates to
reproduction is an adaptation strategy, leading to the partitioning of
reproductive organs, that is, the availability and timely mobilization of
adequate resources from the vegetative plant body to reproductive structures
(Arroyo et al., 2013). Thus far, probably due to reproductive organs'
comparatively minor biomass production and difficulty in being collected,
studies on their decomposition have been limited particularly compared with
those on leaf and other vegetative organs. In this study, we conducted
comprehensive field investigation, pot experiments of litter addition, and
litter-bag experiments to address the following questions.
Should flower litter be considered in the alpine ecosystem's
biogeochemical cycles for their relatively non-negligible biomass production
and/or allocation?
Does flower litter of higher quality and with unique traits have faster
decomposition than leaf litter?
Does the time of litter fall influence soil available nutrients and soil
microbial productivity of alpine meadow ecosystems?
Materials and methods
Study area
The field site is located at the foot of Mt. Kaka, which belongs to the middle section of Minshan
Mountains, eastern Tibetan Plateau (Fig. 1), with a mean annual precipitation
of 720 mm. More than 70 % of precipitation falls in summer from June to
August. Snowfall usually occurs from the end of September to early May next
year. Vegetation presents a typical alpine meadow with numerous and unique
alpine plants. Mosses are abundant and cover most of the ground. The moss
layer is dominated by Polytrichum swartzii and Trematodon acutus c. mull. Vascular plants include species mainly belonging to genus
Kobresia and genus Carex. Other common species are
Festuca spp., Gentiana spp., and Leontopodium spp.
Plant roots in this ecosystem are generally confined to the surface A-horizon
(2–20 cm). A few dwarf shrubs are scattered sporadically in the meadow,
e.g., Rhododendron and Salix. The soil type is dominated by
Mat Cry-gelic Cambisols (i.e., silty loam inceptisol, Chinese Soil Taxonomy
Research Group, 1995).
Plant litter sampling
During the blooming period from the end of May until mid-June and from the
end of July until early August, flower litters of 14 earlier flowering plants
species and 15 later flowering plant species were carefully collected in 2012
at two sites, namely, Mt. Kaka (103∘42′ E, 32∘59′ N;
3500–3900 m a.s.l.) and Bow Ridge Mountain (103∘42′ E,
33∘1′ N; 3600–3850 m a.s.l.). In the study, four litter traps
were placed under the crown of each individual shrub in different communities
(five to eight individuals were chosen for the placement of litter traps),
which were processed and modified based on the litterfall monitoring protocol
(Muller-Landau and Wright, 2010). The litter trap was composed of one cloth
bag and four support legs. A window screen (with a mesh size of 0.8 mm) was
used to seize the cloth bag. Its size was about 50 cm deep and 25 cm long.
Four legs (made with an 80 cm PVC pipe) were tied with a cloth bag and
frame. The frame of the opening was made of iron wire with a 3 mm diameter.
After inserting it into the soil under the shrub's crown, the plant litter
was collected twice per week, which was later sorted as flower litter and
other types during the blooming period. Given the small size of herbaceous
individuals, flowers were plucked at the end of the flowering phase, and
their mass ratios to aboveground biomass were calculated. Freshly fallen
leaves of different species were collected from the floor of the alpine
meadow (i.e., mixed leaf litters, ca. 3950 m a.s.l.). These species were
tentatively classified into five groups according to Raunkiaer's life-form
system (i.e., chamephyte, geophyte, hemicryptophyte, phanerophyte, and
therophyte). Target species were first decided by visual observation. For
herbaceous species, their dominances were determined using quadrat methods.
Each quadrat (1 m × 1 m) was spaced at least 2 m apart from
another along the transect for recording community composition (totaling 10
quadrats along one transect and three transects at each site). Weighted means
of frequency and biomass of target species were sorted and used to assess
their dominances. For shrubs, the line-point intercept method was conducted
to calculate targeted species' frequency, height, and cover, which are
represented by a “hit” (three transects at each site; a 20 m rope with ca.
1 cm diameter or a measuring tape was used), whose weighted means were
sorted to determine the dominant species (Herrick et al., 2005). We also
consulted an expert who has prior knowledge or research on the dominant
species at the selected sites.
These species were divided into earlier flowering species and later
flowering species groups based on blooming time (Table 1). According to
Raunkiaer's life-form system, earlier flowering species mainly consisted of
hemicryptophyte, geophyte, and phanerophyte, whereas more than half of later
flowering species comprised chamaephyte. Nearly half of the tested species
were dominant or co-dominant in their respective communities. The dry matter
content of flower litters in all of the species was ranked from 10 to 60 %. Mixed leaf litters of alpine meadows
were sampled on Mt. Kaka (3950 m a.s.l.), and leaf litters of 13 dominant species were collected to
compare their chemical properties with flower litters. Both types of litters
were first spread on blotting paper for air drying. A small portion of each
litter was further dried in an oven for 48 h to calculate dry matter
content.
Experimental design
Polyvinyl chloride (PVC) pots (15 cm deep, 20 cm diameter at the top, and
12 cm diameter at the bottom) were filled with 2 kg of soils, which were
collected in the autumn of 2011. The collected soil samples were stored at
4 ∘C. The samples were sieved through a 2 mm mesh and then mixed
thoroughly. The soil surface of each treatment was added with 5 g of flower
litters or mixed litters (calculated as dry weight) on 21 June (14 species,
earlier flowering plants) and 11 August 2012 (15 species, later flowering
plants). The surface was covered with a thin layer of soil to avoid being
blown by wind. Another two additional treatments were conducted without
litter addition (control) and with mixed leaf litter addition, respectively.
In total, the pot experiment consisted of 33 treatments with 3 replicates,
with a total number of 99 pots. All of the pots were carefully buried 12 cm
deep into the field to maintain the same soil temperature in the experimental
field. The pots were randomly distributed, and their top edges were
approximately 3 cm above the ground to prevent runoff from outside. All of
the pots were rearranged every week to create a similar microclimate. After
50 days, each soil sample was collected from three points of each pot in the
center and then mixed to avoid the boundary layer effect. Each soil sample
from different PVC pots was mixed evenly by sieving through a 2 mm mesh. The
samples were stored and marked separately in an ice box prior to chemical
determination.
General description of flower litters.
Life
Size of
Dominant
Color
Dry matter
form
inflorescence (cm)
(Y/N)
content (%)
Caragana jubata
C
1–1.5
N
white
29.81
Primula orbicularis
H
1.5
Y
yellow
23.29
Potentilla anserina
G
1–1.8
Y
yellow
51.9
Rhododendron capitatum
P
2–3
Y
purple
32.84
Viola rockiana
H
1
N
yellow
25.22
Myricaria squamosa
P
0.5–1
N
pink
30.95
Potentilla saundersiana
G
1–1.4
N
yellow
54.01
Taraxacum lugubre
H
3–4
Y
yellow
14.97
Aster tongolensis
H
4–5
N
blue
28.72
Cardamine tangutorum
G
0.8–1.5
N
lavender
13.08
Spiraea alpina
P
0.5–0.7
Y
fallow
32.58
Caltha scaposa
H
3–4
Y
yellow
30.43
Rhododendron przewalskii
P
4–5
Y
pink
33.33
Meconopsis integrifolia
H/T
5–7
N
yellow
21.79
Stellera chamaejasme
C
0.5
N
red
28.11
Potentilla fruticosa
P
2–3
Y
yellow
30.43
Meconopsis punicea
H/A
5–8
N
red
33.57
Meconopsis violacea
H
4–6
N
purple
35.70
Sibiraea angustata
P
0.8
Y
white
29.50
Polygonum macrophyllum
H
0.2
Y
pink
21.79
Pedicularis megalochila
C
0.8–1
N
red
33.57
Ligularia virgaurea
C
1.5
N
yellow
16.78
Pilose Asiabell
C
2–2.5
N
pale/green
22.26
Oxytropis ochrocephala
C
1
N
fallow
28.72
Pedicularis longiflora
C
0.8
N
yellow
28.11
Hedysarum vicioides
C
1
N
pink
30.02
Gentiana sino-ornata
C
3–5
Y
purple
44.10
Leontopodium sinense
C
0.2–0.5
Y
white
56.92
Cremanthodium lineare
G
1.2–1.7
Y
yellow
48.93
Note: C, H, G, P, and T represent chamaephyte, hemicryptophyte, geophyte
(one of the subdivided groups in cryptophytes), phanerophyte, and
thermophile, respectively. Y and N indicate whether the species is dominant
or not in the community. The first 14 species are earlier flowering species,
and the other 15 species are later flowering species.
Decomposition rate
A litter bag with a size of 14 cm × 20 cm was used to determine
the decomposition rate of different plant litters. The bag was double-faced
and made from nylon net material with above (4.5 mm × 4.5 mm mesh)
and below (0.8 mm × 0.8 mm mesh) layers. The above layer with a
larger mesh size allowed free access for most micro-arthropods, which
dominate the soil fauna of alpine meadow on the eastern Tibetan Plateau,
whereas the below layer with a smaller mesh size can reduce litter spillage
from the litter bags in the process. As representative species, flower
litters of Rhododendron przewalskii and Meconopsis integrifolia and mixed litter were packed into litter bags with the edges
sealed on 21 June 2012. The litter-bag experiment was conducted to compare
the decomposition rate of flower litters and mixed litter. Each treatment had
eight replicates. After 7 weeks (8 August 2012), the debris or mud was
removed outside the litter bags carefully, then litters were taken outside,
sank into a small water basin for a short time, and sorted out clay and
litter through a 0.5 mm mesh filter. Lastly, the remaining litters were
dried in an oven for 48 h (65 ∘C) and measured the weight on the
balance (accuracy 0.001 g) for decomposition calculation. Litter
decomposition rates can be determined by the following equation:
DR=(P-R)/P× 100,
where DR is the decomposition rate, P is primary litter mass in the litter bags,
and R refers to residue litter before determining percentage mass loss.
Chemistry determination of soil and plant
For soil samples, total dissolved N (TN) contents were determined using
unsieved fresh moist soil subsamples. Soil subsamples were extracted using
2 M KCl and shaken for 1 h at room temperature (20 ∘C), with a
soil-to-solution ratio of 1:5 (weight / volume). The extracted solution
was filtered through filter paper before further determination (Jones et al.,
2004). NH4+–N and NO3-–N were analyzed with the indophenol
blue colorimetric (Sah, 1994) and ultraviolet spectrophotometry methods
(Norman et al., 1985), respectively. Dissolved organic nitrogen (DON) was
calculated by subtracting dissolved inorganic N (NH4+–N and
NO3-–N) from TN. Soil solutions were extracted by centrifugal
drainage, whereas the exchangeable pool was extracted with 2 M KCl by using
the methods reported by Jones et al. (2004). Total phosphorus (TP) consists
of phosphorus mineral and organic phosphorous compounds in the soil, which
can be converted into the dissolved orthophosphate. Available phosphorous
(A-P) is the fragments in soil that can be absorbed by plants, which consist
of water-soluble phosphorus, some adsorbed phosphorus, organic phosphorus,
and precipitated phosphorus in certain soil types. Chemically, A-P is defined
as the phosphorus and phosphate in a soil solution that can be isotope
exchanged with 32P or can be easily extracted by some chemical reagents.
TP and A-P in soils were estimated by extraction with 0.5 M sodium hydroxide
sodium carbonate solution (Dalal, 1973). Microbial biomass carbon (MBC) and
microbial biomass nitrogen (MBN) contents were determined through the
chloroform–fumigation direct-extraction technique. Correction factors of
0.54 for N and 0.45 for C were used to convert the chloroform labile N and C
to microbial N and C (Brookes et al., 1985). For plant samples, the contents
of C and N were determined by dry combustion with a CHNS auto-analyzer system
(Elementar Analysen Systeme, Hanau, Germany; Brodowski et al.,
2006). The content of P was obtained
colorimetrically by the chloro-molybdophosphoric blue color method after wet
digestion in a mixture of HNO3, H2SO4, and HClO4 solution
(Institute of Soil Academia Sinica, 1978). Lignin and cellulose were estimated by the method described by
Melillo et al. (1989).
Production of flower litters and biomass allocation of
representative dominant species. (a) Production of flower litters
and non-flower litters of shrubs (phaenerophyte, n= 20) per unit area
(m2), and (b) floral biomasses and their allocation in the
aboveground biomass.
Results
Flower litter production of dominant species and their biomass
allocation
Among 13 dominant species, the flower litters of phenerophyte plants, whose
flower litters are comparable with non-flower litters, were calculated
through comparison with non-flower litters in the process of flower litter
collection (Fig. 2a). The dry weights of flower litters were
10–40 g m-2, whereas their non-flower litters were only
5–25 g m-2. Although neither of the flower litters of S. angustata or R. capitatum were significantly different compared
with their non-flower litters (P > 0.05), the difference
between the two remained noticeable, whose values were
28.03 ± 3.56 g m-2 vs. 13.21 ± 1.49 g m-2 for
R. capitatum and 19.58 ± 3.50 g m-2 vs.
12.95 ± 0.61 g m-2 for S. angustata. The production of
flower litters was higher than that of non-flower litters. The other three
species significantly produced more flower litters than non-flower litters
(R. przewalskii: F= 15.76, P < 0.001; P. fruticosa: F= 4.76, P < 0.05; S. alpine: F= 10.18, P < 0.01). The flower litters of the eight herbaceous
species were compared with their individual aboveground biomass (Fig. 2b),
which ranged from 10 to nearly 40 %. This
finding indicated that flower litter should be considered to determine the
effect of plants on the soil nutrition pool during the growing season.
Chemical composition and their comparison between flower and leaf
litters. Whiskers refer to quantiles for comparable data settings. Asterisks
(*) represent the distribution of extreme outliers. M= mean and N,
which indicates the data/sample number, are analyzed and processed by one-way
ANOVA (at the P= 0.05 level).
Comparison of chemical properties between flower and leaf
litters
Total C content was not significantly different between flower and leaf
litters (Fig. 3a, F= 1.80, P= 0.199). However, the levels of
cellulose, lignin, and structure C of leaf litter were significantly higher
than those of flower litter (F= 6.74, P < 0.05; F= 5.77,
P < 0.05; F= 10.99, P < 0.01). Hence, flower
litter probably contains more non-structure C than leaf litter.
Both N and P contents of flower litters were significantly higher than those
of leaf litters (Fig. 3b). N in flower litters was nearly doubled to that of
leaf litter (23.17 ± 1.52, 11.87 ± 0.77; F= 45.70, P < 0.001). More than twice the amount of P was also present in
flower litters (2.95 ± 0.25 g kg-1) compared with that in leaf
litters (1.12 ± 0.12 g kg-1; F= 43.87, P < 0.001).
For the implication of the ratio of different chemical properties, C / N,
N / P, and lignin / N were determined to compare flower and leaf
litters. All three indicators of leaf litter were significantly higher than
those of flower litters (Fig. 3c). As parameters used to demonstrate
decomposition rate, C / N and lignin / N of leaf litter were nearly
double those of flower litter (39.27 ± 4.16, 19.80 ± 1.39, F= 37.78, P < 0.001; 21.09 ± 2.25, 12.79 ± 1.15, F= 7.91, P < 0.01). Furthermore, the N / P of flower litter
was significantly higher than that of leaf litter (8.42 ± 0.42,
11.60 ± 0.56; F= 20.62, P < 0.001). These findings
indicated that flower litter can supply more P per unit N than leaf litter.
Assessing the effects of flower litter on soil N pool and P pool
Earlier flowering species exerted positive effects on soil TN,
NO3-–N, and NH4+–N (Fig. 4a), with the addition of their
flower litters according to their size of α values. Most parameters
were higher than 0, which indicated that N2 > N1.
Flower litter increased the soil N pool. All of the minimum α values
of the five indices were also higher than 0 (Table 2, 0.42–1.29), which
indicated that flower litter addition significantly increased the soil N
pool, including different fragments (P < 0.001). Among the later
flowering species, except G. sino-ornata and L. sinense,
soil N indices were significantly improved with flower litter addition, as
demonstrated through α values higher than 0 (Fig. 4b, Table 2). Later
flowering species differed from earlier flowering species, with minimum
α values lower than 0, which resulted from the exceptions of
G. sino-ornata and L. sinense. However, all of the mean
α values were higher than 0, which presented general results after
flower litter addition (0.36–1.49); the soil N pool was significantly
enhanced only after 50 days (P < 0.001). Interactions between
flowering time and litter addition for NO3-–N and NH4+–N
were significant (F= 5.043, P < 0.05; F= 7.947, P < 0.01; F= 24.143, P < 0.05, respectively) but
not for TN (F= 0.470, P= 0.496). Different flowering times
significantly affected NO3-–N, and NH4+–N (Table 3, P < 0.01) but did not significantly influence TN (F= 2.80, P= 0.10). As illustrated in Fig. 4, litter addition had significant effects
on all of the N fragments, which was in accordance with the results in
Table 3. The interaction of flowering time and litter addition exerted
similar effects on the soil N pool as well as its N fragments with flowering
time solely.
Variation in the soil N pool and P pool after addition of flower
litters, (a) earlier flowering species, and (b) later
flowering species. Scatters represent α mean values of different
indexes. Significant differences of deviations from the 0 lines are tested at
the P= 0.05 level (n= 3). TN, NO3–N, NH4–N, TP, and A-P
represent total nitrogen, nitrate nitrogen, ammonium nitrogen, total
phosphorus, and available phosphorus, respectively.
Flower litters exerted different effects on soil TP and A-P. Soil TP
increased in treatment with early flowering litters (Fig. 4a, Table 2, F= 8.498, P= 0.007) but not in later flowering litters (Fig. 4b, Table 2,
F= 0.97, P= 0.33). The minimum α values were lower than 0 (-0.04 and
-0.20, respectively). However, the A-P of both litter treatments was
significantly positively stimulated (F= 47.39, P < 0.001; F= 68.82,
P < 0.001), whose α values were both higher than 0 (0.67–0.13 and
0.06–0.37, respectively). Multifactorial analysis indicated that soil TP
was not significantly different between the sample treated with flower
litter and the control (Table 3, F= 1.07, P= 0.37). No significant
interaction was evident between flowering time and litter addition
treatments on soil TP (F= 0.01, P= 0.93). Litter addition treatments alone
only had a marginal significant effect on soil TP (F= 3.17, P= 0.08).
Moreover, both minimum α values were lower than 0, but TP was not
significantly different between treatments with later flowering litters and
control treatment (F= 0.97, P= 0.33), which mainly resulted from
G. sino-ornata, L. sinense, and C. lineare. Nevertheless, A-P increased significantly after flower litter
addition (F= 43.01, P < 0.001), with a significant interaction
between flowering time and litter addition (F= 6.44, P < 0.05).
Effects of flower litter addition on the soil solution N pool and soil MBC and
MBN
The soil solution N pool has been improved noticeably from 31.46 to
47.35 mg g-1 in flower litter treatment compared with the control,
particularly in the fragment of NO3-–N, which has been greatly
increased (from 30.93 to 46.8 mg g-1; Table 4). In mixed leaf litter
treatment, no obvious variations were found after litter decomposition, with
32.4 mg g-1 NO3-–N and 0.45 mg g-1 NH4+–N,
respectively. Notable differences in both MBC and MBN were found between
different treatments. Litter addition increased not only soil microbial
biomass C (102.05, 68.08, and 46.25 mg kg-1 for flower litter, mixed
litter, and control, respectively) and MBN (73.02, 69.29,
67.13 mg kg-1 for flower litter, mixed litter, and control,
respectively) but also their C / N ratios (1.40, 0.98, and 0.69 for
flower litter, mixed litter, and control, respectively).
Comparison of the mean values of the soil solution pool and soil
microbial biomass between litter addition treated (flower litter and mixed
leaf litter) and control.
Treatments
Soil solution N
Soil microbial
pool (mg g-1)
biomass (mg kg-1)
NO3-–N
NH4+–N
MBC
MBN
MBC/MBN
Flower litter
46.8
0.55
102.05
73.02
1.40
Mixed leaf litter
32.4
0.45
68.08
69.29
0.98
Control
30.93
0.53
46.25
67.13
0.69
Comparison of decomposition rate between flower litter and mixed leaf
litter
R. przewalskii and M. integrifolia are two typical plant
species widely distributed and easily collected. Both species were assessed
to compare decomposition rates of their flower litter and mixed leaf litter.
Differences in decomposition rate among the flower litter of two species and
mixed litter were supposed to be significant (Fig. 5, F= 130.34, P < 0.001). The flower litters of R. przewalskii and
M. integrifolia decomposed much faster than mixed leaf litter.
Moreover, within only 50 days, more than 20 % of R. przewalskii
and M. integrifolia flower litters decomposed, whereas the
decomposition rate for mixed leaf litter was approximately 6 % only
(i.e., the former was nearly 3 times faster). Moreover, no significant
differences were evident in the decomposition rates of the flower litter of
R. przewalskii and M. integrifolia (P= 0.371).
Discussion
Plant litter decomposition is a critical step in the formation of SOM,
mineralization of organic nutrients, and C balance in terrestrial ecosystems
(Austin and Ballaré, 2010; Cotrufo et al., 2015). At an early stage of
decomposition, there exists partial correlation between decomposed plant
material and light fraction in the SOM pool at a transitional stage of the
humification process (Leifeld and Kögel-Knabner, 2005). Species-specific
variations in plant phenology can affect production of litter fall, which is
noticeable during the growing season from the aspect of nutrient cycling,
although the peak of litter fall happens in autumn (the Northern Hemisphere).
Thus, the early litter fall of alpine plants during the study period from May
to August can be a potential nutrient source when nutritional demands
increase for rapid growth and development. In particular, the amount of
flower fall in the study area exceeds the leaf fall during the blooming
season. A previous study indicated that reproductive litter production
accounted for < 10 % of the total litter in January–August and
13–26 % in September–December (Sanches et al., 2008), which was mainly
triggered by rainfall variability that directly altered litter production
dynamics and indirectly altered forest floor litter. In addition, the flowers
are more nutritional than the leaves in terms of nutrients necessary for
plant growth (Lee et al., 2011). In this study, summit production of flower
litters booms during special periods for both earlier flowering and later
flowering species. Flower biomass of herbaceous plants accounts for 10 %
to approximately 40 % of total aboveground biomass. Moreover, these
flower litters were produced considerably earlier than other aboveground
litters that dropped at the end of the growing season. Furthermore, flower
litters and non-flower litters (mainly constituted of leaves) of woody plants
were 10–40 and 5–25 g m-2, respectively, which clearly implies that
flower litter can be a comparable decomposition substrate in alpine
ecosystems, even for phenerophyte plants.
Percentage of decomposed dry mass of M. integrifolia and
R. przewalskii in a 50-day litter-bag study. Column
represents the mean, and bar indicates the standard error (n= 8).
Different lowercase letters indicate significant differences of the
decomposition rate between litter materials (at the P= 0.05 level).
Variation in the soil nutrition pool with flower litter addition.
The histogram for α values of DIN (a) and A-P
(b) indicates the change between treatments and control.
Litter production and decomposition are controlled by biological and physical
processes, such as the activity and composition of soil and litter fauna and
climate variations (Meentemeyer, 1978; Cornejo et al., 1994; Wieder and
Wright, 1995; Aerts, 1997; Cleveland et al., 2004). An integration of index
or traits has been recommended to indicate the process and rate of litter
decomposition. Generally, tissues with high lignin, polyphenol, and wax
contents and higher lignin / N and C / N ratios exhibit slow
decomposition. Lignin / N and C / N ratios are commonly accepted as
good indicators of decomposition rates under short time frames; however,
there is minimal conclusive evidence that lignin is preferentially preserved
in soils compared with bulk soil over long time periods (Melillo et al.,
1982; Aber et al., 1990; Mikutta et al., 2005; Kleber et al., 2007; Cotrufo et al., 2015).
Moreover, lignin plays a dual role in plant litter decomposition when
photochemical mineralization and abiotic decomposition are considered (Austin
and Ballaré, 2010). Leaf litter with C / N ratios lower than 30 is
known to decompose easily and yield a mull humus type, whereas C / N
ratios above 30 result in N immobilization (Heal et al., 1997) and
decomposition retardation. In this study, flower litter had a significantly
lower C / N ratio (19.80 ± 1.39, less than 30) than leaf litter
(39.27 ± 4.16, more than 30). Structural (lignin, DMC) and chemical (N)
traits are proposed to be better predictors for several high-turnover organs
than structural traits alone (Freschet et al., 2012). Lignin content in
flower litters was significantly less than that in leaf litters
(211.37 ± 8.63 and 237.88 ± 6.89 mg kg-1, respectively; F= 5.77, P= 0.02), similar to cellulose (266.93 ± 4.92 and
283.75 ± 4.21 mg kg-1, respectively; F= 6.74, P= 0.01),
which is one of the major cell wall constituents. All of the results are in
accordance with previous studies. Decomposition rate is negatively correlated
with the concentration of lignin, which is a group of complex aromatic
polymers that serves as a structural barrier impeding microbial access to
labile C compounds (Swift et al., 1979; Taylor et al., 1989; Austin
and Ballaré, 2010; Talbot and Treseder, 2012). Moreover, the absence of significant differences of total C
content in flower litters but with significantly fewer structural
carbohydrates than those in leaf litters indicated that greater
non-structural carbohydrates existed in flower litters. This finding can be
inferred from the contents of lignin and cellulose (Fig. 3a). Hence, flower
litters can promote nutrients that easily complement soil (Parton et al.,
2007) for plants in their entire life
history. Decomposition rates of leaf litters have been considered recently
from their lignin / N or lignin/cellulose (Talbot and Treseder,
2012; Cornwell et al.,
2008). Furthermore, in the present
study, lignin / N was less in flower litters (almost 50 % in leaf
litters, i.e., 12.79 ± 1.15 and 21.09 ± 2.25, respectively),
whereas N / P was higher than that of leaf litters.
A litter-bag experiment on two widely distributed dominant shrubs (R. przewalskii and M. integrifolia) confirmed that the decay rates of
flower litters were significantly faster than those of other litters, which
is in accordance with the fast decomposition of R. pseudoacacia
flower from an experiment performed in Korea (Lee et al., 2010). Flower
litters contained significantly higher N and P contents than leaf litters
(Fig. 3b). Plant litter available to the decomposer community encompasses a
broad range of issues that differ in chemical and physical properties (Swift
et al., 1979). P has been regarded as
essential for a long time, which leads to limited attention to mechanisms
that drive P limitation and their interactions with the N cycle (Vitousek et
al., 2010). In most soils, the
concentration of orthophosphate in solution is low (Richardson et al., 2009).
Although soil generally contains a large amount of total P, only a small
proportion is immediately available for plant uptake from the soil solution.
P is derived mainly from rock weathering and the related biogeochemical
cycle, and ecosystems begin their existence with a fixed complement of P, and
even very small losses cannot be readily replenished (Walker and Syers,
1976). The present study indicated
that decomposition of flower litter can be one of the beneficial sources of
soil A-P in alpine ecosystems. Decomposition rates can be markedly affected
by particle size, surface area, and mass characteristics (Angers and Recous,
1997). In addition, physical toughness (lignin, dry matter content, or C
content) can be a suitable predictor of decomposition across all of the
organs. Nevertheless, the current study regarding the characteristics and
driving mechanism of this source remains at the
first stage. Variation in soil physical–chemical properties, vegetation
types, and microbial activities can significantly affect chemical
compositions and forms as well as the biological availability of soil P
directly or indirectly.
Decay rates of different plant organs reflect the
diversity that fruits decompose faster than leaves, which in turn decompose
faster than woody plant parts (Swift et al., 1979; Kögel–Knabner,
2002). Flower litters decompose
rapidly with higher N and P levels supplied to soil, particularly from
NO3-–N in the soil solution pool (Table 4). The histogram for
α values of DIN and A-P also presented soil available nutrients
positively stimulated by flower litter (Fig. 6) for their values distributed
at an interval greater than 0. The high DOC values in flower litter may
influence N and P in soil through the C substrate supplement for soil
microorganisms to enhance N immobilization. Recent empirical studies noted
that the changing microbial community composition significantly affects
ecosystem processes, such as litter decomposition (Strickland et al.,
2009; Ramirez et al.,
2012). Shifts from bacterial-dominated
to fungal-dominated decomposition happened over short (days to a few months)
periods (Poll et al., 2008; McMahon et
al., 2005). Although the present study
did not present the precise analysis of the microbial community, both MBC and
MBN differed greatly between different treatments (Table 4). Litter addition
increased them obviously, which is evident not only in microbial biomass C
and N, but also in their C / N ratios (1.40, 0.98, and 0.69 for flower
litter, mixed litter, and control, respectively). Flower litter contains more
than twice the MBC (increased from 46.25 to 102.05), and both MBC and MBN
pools increased potentially after flower litter addition. Therefore,
microbial functional groups might be changed for nutrient supplement from
litters or could also be due to their faster turnover or growth, which need
more evidence in further study by directly testing soil microbial community
composition.
Several unexpected species in the experiment reduced soil available
nutrients, probably because their specific chemical properties, which change
as a result of microbial activities and nutrient dynamics (Karmarkar and
Tabatabai, 1991), may negatively
affect soil microorganism biomass or activities (Wardle et al., 1998;
Cipollini et al., 2012). Furthermore,
soil microbial communities can be modified through time in response to
allelopathic plants, with known or potential effects on plant communities
(Cipollini et al., 2012; Inderjit, 2001). Soil carbon generally is divided into pools with varying
intrinsic decomposition rates in turnover models, whose decomposition rates
can be modified and codetermined by interaction between substrates, microbial
actors, and abiotic driving variables. These factors are rationalized by
assuming chemical structure is a primary controller of decomposition (Kleber
et al., 2010). Most of the non-fertilizer N source needed for plant growth is
SOM (Sollins et al., 2007), which consists of organic molecular fragments
with wide-ranging amphiphilicity degrees, intimately contacting mineral
surfaces of variable chemical reactivity and a polar solvent. Mineralization
and nitrification can be subdued by inhibitory compounds from the exudates of
a certain plant species, which come from a negative aspect and mainly result
from suppression of related microbes (Cipollini et al., 2012). In another
positive perspective, considering the “priming effect” once flower
litter is added in moderate treatments causes strong short-term changes in
the turnover of SOM, and nutrient release follows litter decomposition
(Jenkinson et al., 1985; Kuzyakov et al., 2000; Blagodatskaya and Kuzyakov,
2008). Hence, N and P
availability in the soil of alpine ecosystems can be maintained in part by
tissue chemistry favorable to microbial decomposition and release of
nutrients.
Flower litter influences different fractions in soil N and P pools as well as
soil microbial biomass (i.e., MBC and MBN), which provided evidence that
plant species, through tissue chemistry, biomass allocation, and phenology,
affect local soil properties and SOM formation in alpine ecosystems. Soil has
a specific susceptibility to decomposition of biochemical compounds in plant
tissues, on a spectrum from quickly decomposed labile to relatively
recalcitrant. Flower litters have intuitive benefits chemically and
physically for the formation, stabilization, and mineralization process of
SOM. In future studies, major scientific findings and also potential
questions less studied previously should be highlighted, and scientific
obstacles should be considered to further address the stabilization and
destabilization of SOM in this field. In brief, under a changing climate and
a steadily increasing service demand in the alpine ecosystems, it is
essential to understand the mechanisms underlying SOM stabilization.
Furthermore, soil carbon models would benefit from taking flower litters'
decomposition with specific attribution into soil nutrition pools. Flower
litters affect carbon and nutrient cycling and should be incorporated into
SOM pools along with decomposition simultaneously, which should be enhanced
in future studies to better understand the essentiality and fundamentality of
litter decomposition.