Ecology, 95(7), 2014, pp. 1861–1875
! 2014 by the Ecological Society of America
Effects of arctic shrub expansion on biophysical vs. biogeochemical
drivers of litter decomposition
JENNIE DEMARCO,1,4 MICHELLE C. MACK,2
AND
M. SYNDONIA BRET-HARTE3
1
Department of Biology, New Mexico State University, Las Cruces, New Mexico 88003 USA
2
Department of Biology, University of Florida, Gainesville, Florida 32611 USA
3
Institute of Arctic Biology, University of Alaska, Fairbanks, Alaska 99775 USA
Abstract. Climate warming in arctic tundra may shift dominant vegetation from
graminoids to deciduous shrubs, whose functional traits could, in turn, alter biotic and
abiotic controls over biogeochemical cycling of carbon (C) and nitrogen (N). We investigated
whether shrub-induced changes in microclimate have stronger effects on litter decomposition
and nutrient release than changes in litter quality and quantity. In arctic tundra near Toolik
Lake, Alaska, USA, we incubated a common substrate in a snow-addition experiment to test
whether snow accumulation around arctic deciduous shrubs altered the environment enough
to increase litter decomposition rates. We compared the influence of litter quality on the rate
of litter and N loss by decomposing litter from four different plant functional types in a
common site. We used aboveground net primary production values and estimated decay
constant (k) values from our decomposition experiments to calculate community-weighted
mass loss for each site. Snow addition had no effect on decomposition of the common
substrate, and the site with the highest abundance of shrubs had the lowest decomposition
rates. Species varied in their decomposition rates, with species from the same functional type
not always following similar patterns. Community-weighted mass loss was 1.5 times greater in
the high shrub site, and only slightly decreased when adjusted for soil environment, suggesting
that litter quality and quantity are the primary drivers of community decomposition. Our
findings suggest that on a short time scale, the changes in soil environment associated with
snow trapping by shrubs are unlikely to influence litter nutrient turnover enough to drive
positive snow–shrub feedbacks. The mechanisms driving shrub expansion are more likely to
do with shrub–litter feedbacks, where the higher growth rates and N uptake by shrubs allows
them to produce more leaves, resulting in a larger litter N pool and faster internal cycling of
nutrients.
Key words: Arctic; climate change; deciduous shrubs; decomposition; litter quality; snow manipulation.
INTRODUCTION
Temperatures in the Arctic region are warming at
twice the rate of the rest of the globe (Hassol 2004,
Serreze and Francis 2006, Kaufman et al. 2009, Screen
and Simmonds 2010), altering the structure and function
of ecosystems (Chapin et al. 1995, Chapin and Shaver
1996, Shaver et al. 2000, Mack et al. 2004) by causing a
shift in species composition from one previously
dominated by graminoids to an increase in the
abundance and expansion of shrubs (Jia and Epstein
2003, Stowe et al. 2004, Tape et al. 2006, Jia et al. 2009,
Forbes et al. 2010, Elmendorf et al. 2012). Increasing
shrub dominance will increase plant productivity and
biomass, resulting in more uptake of atmospheric CO2
(Cahoon et al. 2012) and the storage of more C
aboveground in woody tissue or belowground in
rhizomes (Shaver and Chapin 1991), roots, and ectoManuscript received 3 December 2013; accepted 17
December 2013; final version received 8 January 2014.
Corresponding Editor: R. W. Ruess.
4 E-mail: [email protected]
mycorrhizal fungi (Clemmensen et al. 2006). In addition, more woody stems associated with shrubs could
mean more C stored in coarse woody debris and stems,
which decompose slowly (Hobbie 1996). Taken together, these effects of shrub expansion have been identified
as a negative carbon cycling feedback to climate
warming. Biophysical feedbacks to regional climate,
however, may be in the opposite direction. Shrublands
have a lower albedo than tundra, which can lead to an
increase in absorbed solar radiation during the snowfree period, resulting in a positive feedback to regional
warming (Chapin et al. 2005). In addition, deciduous
shrubs in fertilized tundra cycled C and N faster than in
unfertilized tundra, resulting in a net loss of deep soil C
(Mack et al. 2004). Soils in naturally occurring shrub
tundra similarly store less C than those in graminoid
tundra (Ping et al. 1997). Thus, net positive feedbacks to
warming are possible if increased shrub cover alters
ecosystem structure and function so that soil C stocks
decrease. Since the Arctic stores 20–30% of the total
amount of terrestrial soil-bound C (McGuire et al. 2009)
and woody vegetation is predicted to increase by as
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JENNIE DEMARCO ET AL.
much as 52% by 2050 (Pearson et al. 2013), it is
important to better understand the mechanisms behind
these potential feedbacks to climate change.
The mechanisms that drive shrub expansion are not
well understood. It has been hypothesized that shrubs
may enhance their dominance and growth through
increased nutrient availability associated with shrubinduced changes in the abiotic environment. The snow–
shrub hypothesis suggests that taller and more abundant
shrubs accumulated greater snow depth due to greater
retention of snowfall (e.g., less snow lost to wind events)
and trapping of wind-distributed snow than tundra
areas with fewer shrubs (Sturm et al. 2001, 2005,
Pomeroy et al. 2006). This deeper snow cover insulates
the soil, maintaining warmer temperatures through the
winter. By altering the abiotic environment through
trapping snow and maintaining warmer soil temperatures, shrubs can influence the biogeochemical processes
that drive nutrient cycling. Decomposition of litter is the
major pathway by which nutrients are recycled and
made available for plant uptake in these nutrient-limited
systems. Changes in the environment can influence rates
of litter decomposition. However, it has not been
directly tested whether the amount of snow trapped by
deciduous shrubs alters the environment enough to
stimulate litter decomposition and increase litter nutrient release.
Previous studies have shown that snow acts as an
insulator that can increase soil temperature (Brooks et
al. 1996, 1998, Grogan and Jonasson 2003, Schimel et al.
2004, Wahren et al. 2005) and the availability of water
to soil microorganisms (Romanovsky and Osterkamp
2000, Mikan et al. 2002), potentially regulating the rate
at which microbes and fungi can break down litter over
the winter. Indeed, litter decomposition has been found
to occur in the winter and under snow (Stark 1972,
Hobbie and Chapin 1996, McLaren and Turkington
2010, Saccone et al. 2013), and to be higher in areas that
have deeper snow cover (Baptist et al. 2010, Saccone et
al. 2013). If faster turnover of litter N results in an
increase in plant-available N, then snow accumulation
by shrubs could indirectly influence N availability by
maintaining warmer soil temperatures in fall and winter,
and allowing a longer window for microbial breakdown
of litter substrates, increased N release, and higher rates
of N supply to plants, resulting in a positive plant–soil
feedback that promotes further shrub expansion.
Shrubs may also have effects on decomposition and N
release that are independent and opposite of their effects
on winter soil temperatures. For example, in moist
acidic tundra, the deciduous shrub Betula nana allocates
79% of its total annual aboveground biomass to new
and old stems (Shaver et al. 2001) that decompose three
times slower than leaves and one to eight times slower
than leaves and stems from graminoids and evergreen
shrubs (Hobbie 1996). A compositional shift to more
deciduous shrub dominance may thus alter nutrient
turnover through biotic controls by a shift toward larger
Ecology, Vol. 95, No. 7
inputs of slowly decomposing litter (Hobbie 1992,
Buckeridge et al. 2010). By slowing nutrient turnover
and decreasing plant-available N, this litter could drive a
negative plant–soil feedback that would slow shrub
expansion.
These two competing mechanisms lead to the question
of whether changes in microclimate or changes in litter
quality and quantity have stronger effects on litter
decomposition and nutrient release. This question has
been studied in mesic (Hector et al. 2000, Knops et al.
2001, Scherer-Lorenzen 2008) and dry (McLaren and
Turkington 2010) grasslands and temperate forests
(Hobbie et al. 2006, Vivanco and Austin 2008), but
has not been as well studied in Arctic tundra systems,
particularly in the context of shrub expansion. Will
positive or negative plant–soil feedbacks prevail as
shrubs expand in the Arctic? Although there is much
evidence to suggest that shrubs can influence their
environment to alter key biogeochemical processes that
control plant nutrient supply, there have been no
published studies to date that have directly tested the
effect of added snow (at the depth that would be trapped
by shrubs) or increased shrub cover on litter decomposition. In addition, within the Alaskan Arctic, most
decomposition studies have largely occurred in graminoid-dominated moist acidic or nonacidic tundra. We
know relatively little about how the environment of
shrub tundra may influence litter decomposition.
The goal of our study was to understand the relative
importance of mechanisms through which arctic deciduous shrubs affect litter decomposition and N dynamics.
Our objectives were threefold: (1) to test whether snow
addition, at a rate realistic for increasing shrub
abundance, altered the environment for decomposition
enough to stimulate rates of litter mass and N loss, (2) to
compare how litter C quality and the relative availability
of C to N influenced the rate of litter mass and N loss,
and (3) to better understand how the changes in plant
species composition associated with the shift from
graminoid- to shrub-dominated tundra influence community decomposition. We hypothesized that across
three sites that represented a gradient in shrub
abundance and height, (1) the addition of snow would
slow temperature decline in the winter and lead to faster
decomposition and net N release from litter, (2)
decomposition and net N release would covary positively with lignin : N ratios, and (3) community-weighted
mass loss would be lowest in the shrub-dominated sites,
because of the greater abundance of woody litter.
To test our hypotheses, we measured litter quality,
quantity, decomposition, and net N release across three
plant communities that represented natural variation in
shrub abundance across the landscape (DeMarco et al.
2011). In each community, we manipulated snow depth
using snow fences. To test for the effects of site and
snow on decomposition and net N release, we decomposed a common substrate in all sites and treatments.
We decomposed litters from multiple plant functional
July 2014
SHRUB EXPANSION AND DECOMPOSITION
groups in a common environment, hereafter referred to
as a common garden, to control for differences in
microclimate and directly test the effect of litter quality
on decomposition. Finally, we combined these studies
with aboveground net primary production estimates to
calculate community-weighted mass loss rates.
METHODS
Study area
All sites are located near Toolik Field Station at the
Arctic Long Term Ecological Research (LTER) site
(68838 0 N, 149838 0 W, elevation 760 m) in the foothills
region on the North Slope of the Brooks Range, Alaska,
USA. This area is a younger landscape glaciated during
the late Pleistocene. It includes large areas of the Itkillik
I (deglaciated ;60 000 yr) and Itkillik II (deglaciated
;10 000 yr) glacial drifts (Hamilton 1986). The entire
foothills region of the Brooks Range is treeless and
underlain by continuous permafrost, 250–300 m thick
(Osterkamp and Payne 1981). Mean annual air temperature is around !108C, with average summer temperatures from 7–128C. Annual precipitation is 318 mm, with
43% falling as snow (data available online).5 Average
snow depth is 50 cm, although snow distribution can be
variable due to redistribution by wind. Snowmelt occurs
in early May.
In the fall of 2005, three sites were selected for the
snow manipulation experiment that varied primarily in
deciduous shrub abundance, hereafter referred to as low,
medium, and high shrub sites. These sites are described
in detail in DeMarco et al. (2011). In short, sites were
chosen to have similar state factors (climate, parent
material, time since deglaciation) but varied in the
abundance of deciduous shrubs (Jenny 1994). The same
species of deciduous shrubs (Betula nana and Salix
pulchra) are found at all three sites (except S.
richardsonii, which is found only at the medium shrub
site). However, percent cover of deciduous shrubs
increases from 15% to 94%, and their canopy height
increases from 4 cm to 50 cm across sites. All sites are
within 1 km of each other, and have similar parent
material, time since last glaciation (Itkillik I, 60 000 yrs),
and regional climate, although microclimates vary
across sites due to differences in slope and aspect.
Elevation changes from about 764 m at the low shrub
site to 741 m at the medium and high shrub sites.
Our low shrub site was located in moist acidic tussock
tundra, where the vegetation consists of approximately
equal biomass of graminoids (Eriophorum vaginatum
and Carex bigelowii ), dwarf deciduous shrubs (B. nana,
Vaccinium uliginosum, and S. pulchra), evergreen shrubs
(Ledum palustre ssp. decumbens and V. vitis-idea), and
mosses (Hylocomium splendens, Aulacomnium turgidum,
Dicranum spp., and Sphagnum spp.; Shaver and Chapin
1991). Our medium shrub site was located in riparian
5
http://ecosystems.mbl.edu/ARC
1863
tundra where the deciduous shrubs are of intermediate
size. The vegetation consists of graminoids (primarily C.
bigelowii ), deciduous shrubs (B. nana, V. uliginosum, S.
pulchra, and S. richardsonii ), and mosses (H. splendens
and Dicranum spp.). Our high shrub site was located in
riparian tundra dominated by tall deciduous shrubs and
has predominantly deciduous shrubs (B. nana, S.
pulchra, and some Potentilla fruticosa), with some
evergreen or wintergreen shrubs (V. vitis-idaea and
Linnaea borealis), forbs (Polygonum bistorta, Petasites
frigidus, Stellaria longipes, Valeriana capitata, and
Artemisia alaskana), graminoids (Poa arctica, C. bigelowii, and Calamagrostis canadensis), and mosses (Sphagnum spp. and H. splendens).
Snow manipulation
To determine the influence of increased snow depth
on litter decomposition, snow fences that represented
maximum regional shrub height (1.5 m high) were set up
in the fall of 2005 at all three sites to manipulate snow
depth (DeMarco et al. 2011). Two treatments, control
(ambient snow) and drift (manipulated snow), were set
up at each site. For all sites, subplots on the drift side of
the fences were located in the zone of maximum snow
accumulation, which was relatively uniform. Within
each treatment, 18 2 3 10 m plots, with 1-m buffer strips
between, were established. For this study, six plots per
treatment (n ¼ 6) were randomly assigned to measure
litter decomposition. Remaining plots were used for
other experiments not described here.
Soil temperature at 5 cm within the organic layer was
measured continuously (1–3 h intervals) from July 2006–
May 2009 in each study plot (n ¼ 3–4 plots within each
treatment and site) using I-button temperature data
loggers (IButtonLink, East Troy, Wisconsin, USA).
Mean daily soil temperatures were calculated for all
plots within each treatment and site for each year. Mean
growing season and winter soil temperatures were
calculated from the mean daily temperatures from each
plot within each treatment and site. I-buttons were not
always installed or removed on the same day; therefore
we only analyzed data from days in which we had data
for all sites and treatments. The growing season included
measurements taken from 1 July to 1 August of that
year and included years 2006, 2007, and 2008. Winter
growing season includes measurements taken from 1
September through 1 May of the following year and
includes winters from 2006–2007, 2007–2008, and 2008–
2009.
The snow fences produced snow packs in treatment
plots that were, on average, 87, 96, and 104 cm deeper
than ambient snow depth for the low, medium, and high
shrub sites, respectively. Snow addition increased
average winter soil temperatures by 38C and summer
soil temperatures by 28C in the high shrub site; the
medium and low shrub sites showed similar trends,
although the differences between treatments were
smaller in magnitude (DeMarco et al. 2011).
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JENNIE DEMARCO ET AL.
Common substrate experiment
To directly test the effect of microclimate and snow
addition on litter decomposition rates, we incubated the
senesced leaves from a common substrate, B. neoalaskana (Sarg.), in the ambient and snow-manipulated
plots across all three sites. Senesced leaves were collected
from trees growing near Fairbanks, Alaska, USA.
Leaves were still attached to the trees, but the petiole
had already started to abscise. This common substrate
was used because the large leaf size and relative
abundance allowed us to collect enough material for
our study. Leaves were air dried, well mixed, and then
subsampled for litter bags. One gram of leaves was sewn
into 2-mm mesh bags, 8 3 8 cm in size. Litter bags were
incubated beneath the live moss and litter layer starting
in early June of 2006. The moss and litter in this system
are well mixed, so bags were inserted in this layer. Four
identical bags were strung together for four separate
annual harvests. Bags were placed in six treatment plots
in each site, with three sub-replicates within each plot.
Bags were removed in July of 2007, 2008, and 2009, and
were kept frozen until they could be processed.
At time of processing, bags were thawed and then
gently rinsed with deionized (DI) water to remove soil
and loose litter attached to the outside of the bag. All
original leaf litter was removed, dried at 458C for a
minimum of 48 h, and weighed. To determine the
percentage of C and N of the litter, samples were ground
to a fine powder on a Wiley mill, with a no. 40 mesh
screen, and then analyzed using an ECS 4010 elemental
analyzer (Costech Analytical, Valencia, California,
USA). Percentage of initial mass remaining was
calculated by dividing the incubated mass by the initial
mass and multiplying by 100. Percent of initial C (ICR)
and initial N remaining (INR) was calculated by the
following equations:
ICR ¼
INR ¼
ðt1mass 3 t1Carbon Þ
3 100;
ðt0mass 3 t0Carbon Þ
ðt1mass 3 t1Nitrogen Þ
3 100:
ðt0mass 3 t0Nitrogen Þ
Common garden experiment
To compare differences in litter decomposition rates
among species, we incubated senesced leaf litter from 10
vascular plant and three moss species, and stem litter
from four shrub species collected across eight sites
located in the Arctic Foothills region on the North Slope
of the Brooks Range (Table 1). Three of the sites were
the control plots at the low, medium, and high shrub
sites. Four sites were located in alder-dominated (Alnus
viridis spp. fruticosa) tall deciduous shrub tundra; two
near the Sagavanirktok River and two along the Dalton
Highway, ;32 km north of Toolik Field Station. Alders
are one of the deciduous shrubs that have been
documented as expanding in many arctic regions. One
Ecology, Vol. 95, No. 7
site was in the control plots in a previously set up
experiment in moist acidic tundra (referred to as species
removal; Bret-Harte et al. 2008). All litter was collected
and processed using the same methods as described in
the common substrate experiment, except that only 1.6mm mesh bags, 4 3 8 cm in area, were used, and leaf
samples were replicated six times, while stem litter was
replicated three times. For the litter collected at the
species removal site, litter bags were installed in July of
2003 and removed in July of 2004, 2005, 2007, and 2008.
Litter collected from the other sites was installed in the
field as litter bags in early June of 2006 and replicate
bags were removed in July of 2007, 2008, and 2009 and
processed as described previously.
Calculations
The exponential decay constant, k, was determined by
assuming a single exponential decay model (Olson
1963): Mt ¼ M0e!kt, where Mt is litter mass at time t,
and M0 is initial mass. The slope of the regression of
proportion of initial mass remaining against time was
used to determine the decay constant for each substrate
at each site.
Initial litter quality
A subsample of each leaf and stem collection was
analyzed for percentage of C, percentage of N, and C
quality to determine the quality of the litter substrates
prior to decomposition. Percentage of C and N was
determined from samples that had been ground to a fine
powder on a Wiley mill, with a no. 40 mesh screen, and
then analyzed using an ECS 4010 elemental analyzer. C
quality measurements were carried out on an ANKOM
fiber analyzer (Ankom Technology, Macedon, New
York, USA) and included determination of (1) soluble
cell contents (carbohydrates, lipids, pectin, starch, and
soluble protein), (2) hemicelluloses plus bound proteins,
(3) cellulose, and (4) lignin plus other recalcitrants
(Ryan et al. 1990).
Community-weighted mass loss
To separate out the effect of changes in species
composition vs. changes in the microenvironment
associated with shrub expansion on community level
decomposition, we used measured aboveground net
primary productivity (ANPP) and k values to calculate
a community-level mass loss for each of the three shrub
sites (Appendix A: Table A1). We followed the
procedure outlined in Hobbie and Gough (2004). We
harvested biomass in July of 2007 from the control
treatments of the three snow fence sites (DeMarco et al.
2011). Total ANPP (g%m!2%yr!1) per plot (n ¼ 8
replicates per site) was calculated by summing new
apical biomass (g/m2) for that year, i.e., of each tissue
type from each species found within that plot. Our
ANPP calculations were only for vascular plants and
thus do not include mosses, lichens, or belowground
parts such as rhizomes or roots. We estimated the
July 2014
SHRUB EXPANSION AND DECOMPOSITION
1865
TABLE 1. List of all the species decomposed in the common garden site.
Group
Graminoids
Deciduous shrubs
Evergreen shrubs
Mosses
Species
Carex bigelowii, Eriophorum vaginatum
Alnus crispa viridis spp. fruiticosa, Betula nana, Betula neoalaskana
(Sarg.), Rubus chamaemorus, Salix pulchra, Vaccinium uliginosum
Ledum decumbens, Vaccinium vitis-idaea
Aulacomnium turgidum, Hylocomium splendens, Sphagnum spp.
contribution of secondary stems to ANPP by multiplying the ANPP of species likely to produce woody tissue
by a proportion determined by Bret-Harte et al. (2002).
These were 0.158, 0.181, and 0.079 for B. nana, S.
pulchra, and L. palustre ssp. decumbens, respectively. We
followed the methods outlined in Hobbie and Gough
(2004) and assumed that C. tetragona and V. uligonosum
resembled L. palustre in their proportional secondary
growth, and S. reticulate, S. glauca, and S. richardsonii
all resembled S. pulchra in their proportional secondary
growth. We also assumed that Andromeda polifolia,
Arctostaphylos alpina, Dryas integrifolia, Empetrum
nigrum, Rubus chamaemorus, V. vitis-idaea, and Linnaea
borealis had negligible secondary growth. We assumed
inflorescences would have similar k values as foliar litter
from the same species and thus used foliar k values for
inflorescences. This method assumes all new production
that year ends up as litter. For deciduous shrub leaves,
this is a valid assumption, as they lose all of their leaves
at the end of the growing season. Evergreen shrubs and
graminoids can retain leaves through the winter, and
thus we may be overestimating litter production for
these species.
The mass loss for each plant species and their parts
(inflorescences, leaves, and stems) was calculated by
multiplying the species’ ANPP by that species’ decay
constant, k, from the common garden experiment to
produce an ANPP-weighted mass loss for each species.
ANPP-weighted mass loss values were summed across
all species in a plot to produce community-weighted
mass loss values (see Appendix A: Table A5 for full list
of ANPP and k values used for each species present).
To determine how differences in site environments
contributed to community mass loss, we adjusted the
community-weighted k values for site-specific differences
in decomposition rates. We divided site mean k values
for the common substrate, B. neoalaskana, by the mean
k value for B. neoalaskana that was decomposed in the
common garden (k ¼ 0.256). The proportional difference
between the k value from the common garden and the
shrub sites was multiplied by the species-specific k values
to correct for differences in decomposition found across
the shrub sites. The adjusted species-specific k values
were summed to obtain a decay k constant for the entire
community. This method assumes that all species and
tissues respond the same way as B. neoalaskana when
decomposed in different environments, and does not
take into account any site by species interactions.
Evidence of site by species interactions are rare, because
most decomposition studies do not conduct full factorial
incubations where all species are incubated in all
environments. In moist acidic and moist nonacidic
tundra, B. nana stem decomposition followed the
opposite pattern between the two sites compared to
the other species decomposed (Hobbie and Gough
2004).
Species decomposed in our common garden contributed 93% of ANPP. For the species that contributed the
remaining 7% of ANPP, we used k values from other
published studies within the region or substituted k
values for species with similar growth forms. For all
forbs we used the decay constant for P. bistorta reported
by Hobbie and Gough (2004), which was decomposed in
moist acidic tundra not far from our common garden
site. For Calamagrostis spp., Poa arctica, and Juncus
spp., we used an average of the k values for Carex spp.
and E. vaginatum (k ¼ 0.18) from our common garden.
For Empetrum spp., L. borealis, and Cassiope spp.
leaves, we used an average of k values from Ledum spp.
and V. vitis-idaea from our common garden. For stems
of these three species, we used an average of k values
from B. nana and S. pulchra stems (k ¼ 0.08).
Statistical analysis
To test our hypotheses about the effects of site and
snow depth on decomposition of the common substrate,
we used two-way ANOVA (JMP 7.0, SAS Institute,
Cary, North Carolina, USA). Relationships between k
and initial litter quality indices for 10 vascular plant
species decomposed in the common garden were tested
using single regression analysis with k as the dependent
variable and the litter quality indices of interest as the
independent variable. The effect of changes in litter
quality/quantity on community mass loss and the effect
of changes in environment on community mass loss were
tested using two separate one-way ANOVAs with site as
the main effect. Tukey’s HSD test was used as a post hoc
test when ANOVAs were significant at P , 0.05. Data
were tested for normality (Shapiro-Wilks), and lntransformed when necessary to achieve homogeneity of
variance. When homogeneity of variance could not be
achieved, data was analyzed using a Kruskal-Wallis
nonparametric test.
RESULTS
Effect of added snow on soil temperature
Snow addition significantly increased winter soil
temperatures by an average of 1–48C across all three
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JENNIE DEMARCO ET AL.
Ecology, Vol. 95, No. 7
TABLE 2. Three-way ANOVA comparing differences in soil temperature between vegetation type, treatment, and year.
Growing season
Winter
Source
df
F
P
df
F
P
Vegetation type
Treatment
Year
Vegetation type 3 Treatment
Vegetation type 3 Year
Treatment 3 Year
Vegetation type 3 Treatment 3 Year
2
1
2
2
4
2
4
12.0
6.0
18.6
0.7
1.2
0.1
0.2
,0.0001
0.02
,0.0001
0.51
0.30
0.88
0.95
2
1
2
2
4
2
4
26.2
24.5
28.4
3.3
2.4
0.1
1.4
,0.0001
,0.0001
,0.0001
0.05
0.07
0.90
0.25
Note: Error degrees of freedom are 49 for growing season and 35 for winter.
sites (Table 2, Fig. 1). There were significant differences
in soil temperature across sites with snow addition and
across years, but there were no significant interactions
among the main effects. Soil temperatures were highly
variable across years. In general, the low shrub site had
both lower summer and winter soil temperatures
compared to the medium and high shrub sites.
Common substrate experiment
After three years, snow addition did not significantly
alter mass remaining of the common substrate (two-way
ANOVA; treatment; F1,36 ¼ 1.4, P ¼ 0.24 [Fig. 2]) but
there was a significant difference across sites (site; F2,36 ¼
13.6, P , 0.001) and a significant interaction between
site and treatment (site 3 treatment; F2,36 ¼ 3.5, P ¼
0.04). When the effects of treatment were tested for each
site independently, treatment was not significant at the
low shrub site (one-way ANOVA; treatment; F1,11 ¼ 1.9,
P ¼ 0.20) and was only marginally significant at the
medium (treatment; F1,11 ¼ 3.4, P ¼ 0.10) and high shrub
sites (treatment; F1,12 ¼ 3.6, P ¼ 0.09). There was a trend
for greater mass loss in the ambient treatment at the low
shrub site, but greater mass loss occurred in the snowaddition treatment at the medium and high shrub sites
(Appendix B: Table B1). Decomposition rates (k), N
remaining, or C remaining were also not significantly
different under snow addition, although there was a
trend for higher k rates and less N and C remaining in
the snow-addition treatment (Figs. 2 and 3). However,
initial N remaining showed a significant site by
treatment effect. When the main effect of treatment
was tested for each site separately, only the low shrub
site showed a significant effect of treatment on N
remaining after three years (treatment; F1,11 ¼ 122.2, P ¼
0.03). In contrast to the effect of snow addition,
decomposition rates and N and C remaining varied
significantly across sites (Fig. 3). The litter decay rate, k,
was highest at the low shrub site, losing 10% and 6%
more mass and 8% and 6% more C than in the medium
and high shrub sites, respectively. Initial N remaining
followed a different pattern, with mineralization of litter
N occurring at the low shrub site only, but immobili-
FIG. 1. Soil temperature at 0–5 cm soil depth (means 6 SE) during the growing season and winter over three years (2006, 2007,
and 2008), with two snow treatments (ambient and snow addition) and across three shrub sites (low, medium, and high). Data
points that share a lowercase letter are not significantly different at the P , 0.05 level.
July 2014
SHRUB EXPANSION AND DECOMPOSITION
1867
FIG. 2. Initial mass, C, N, and proportion of C:N (means 6 SE) remaining of the common substrate (Betula neoalaskana) from
litter bags incubated over three years in the ambient and snow-addition treatments at the low, medium, and high shrub sites.
zation occurring at both the medium and high shrub
sites. The proportion of initial C:N remaining decreased
with an increase in shrub abundance (Fig. 3).
Common garden experiment
After three years of incubation, neither site of origin
nor species had an effect on the percentage of initial
mass remaining or decay rate of leaf and stem litter from
B. nana and S. pulchra, despite significant differences in
their initial litter quality (Appendix C: Tables C1 and
C2).
Initial leaf litter quality significantly differed across
species for all indices measured, although species within
the same growth form were not always similar in their
initial litter quality (Table 3). Deciduous shrubs had up
to two times more N in their leaves than evergreen
shrubs, graminoids, and mosses (species: v2 ¼ 38.3, df ¼
7, P , 0.0001). Evergreen shrubs had the highest
percentage of C in their leaves, followed by deciduous
shrubs (except R. chamaemorus), graminoids, mosses,
and R. chamaemorus (species: v2 ¼ 40.0, df ¼ 7, P ,
0.0001). Evergreen shrubs, mosses, and the graminoid E.
vaginatum had high C:N ratios in their leaves, while the
deciduous shrubs R. chamaemorus and V. uliginosum
had the lowest C:N ratios (species: F9,53 ¼ 72.2, P ,
0.0001). The evergreen shrubs and the deciduous shrub
B. nana had lignin : N ratios that were two to four times
higher than the deciduous shrubs, R. chamaemorus and
V. uliginosum, and the graminoids (species: v2 ¼ 27.1, df
¼ 6, P , 0.0001). Graminoids had 1.5–4 times more
hemicellulose in their leaves, compared to evergreen and
deciduous shrub leaves (species: F6,32 ¼ 90.1, P ,
0.0001). Graminoids also had the highest percentage of
cellulose in their leaves, two to four times more than in
evergreen and deciduous shrubs. B. nana had the least
percentage of cellulose in its leaves compared to all six of
the other species (species: F6,32 ¼ 157.3, P , 0.001). B.
nana also had the highest percentage of lignin, followed
by evergreen shrubs and the deciduous shrubs, R.
chamaemorus and V. uliginosum. Graminoids had the
least amount of lignin in their leaves (species: v2 ¼ 27.4,
df ¼ 6, P , 0.0001).
After five years of decay, R. chamaemorus leaf litter
lost 1.5–6 times more mass than leaf litter from the other
six species of vascular plants and three species of mosses
collected at the same moist acidic tundra site and
decomposed in the same common garden (Fig. 4; IMR
species: v2 ¼ 36.5, df ¼ 7, P , 0.0001). Species within the
same functional group did not always follow the same
pattern in their rates of decomposition (Fig. 4). Leaf
litter from the deciduous shrubs B. nana and V.
uliginosum had decay constants that were similar to that
of the evergreen shrub L. decumbens, and the graminoid
E. vaginatum had a higher decay constant than those of
the evergreen shrub V. vitis-idaea and the graminoid C.
bigelowii. Mosses had the lowest decay constants
compared to the other seven vascular plant species
decomposed in our experiment.
In a comparison with all 10 vascular species decomposed in our common garden, leaf litter decay rates were
1868
JENNIE DEMARCO ET AL.
Ecology, Vol. 95, No. 7
FIG. 3. Decay constants, initial C and N remaining, and the proportion of initial C:N remaining for the common substrate
(means 6 SE), B. neoalaskana, decomposed across three sites (low, medium, and high) and two snow fence treatments (ambient
and snow addition) and calculated for the entire 3-yr decomposition time period. Results of two-way ANOVAs are displayed for
each variable with site and treatment as the main effects and site by treatment (S 3 T ) interaction. In each panel, data points that
share a lowercase letter are not significantly different at the P , 0.05 level.
TABLE 3. Initial litter quality for senesced leaves of seven species of vascular plants collected from plots in a moist acidic tundra
community and incubated in a common site.
Species
N (%)
C (%)
C:N
Graminoids
Carex bigelowii
Eriophorum vaginatum
1.34 6 0.04
0.87 6 0.04
42.39 6 0.20
42.15 6 0.13
31.91c 6 1.14
8.90b 6 2.11
Deciduous shrub
Betula nana
Rubus chamaemorus
Vaccinium uliginosum
1.67 6 0.06
1.90 6 0.04
1.74 6 0.05
44.59 6 0.15
40.53 6 0.57
43.63 6 0.10
26.94c 6 0.93
21.32d 6 0.38
25.15c,d 6 0.79
Evergreen shrub
Ledum decumbens
Vaccinium vitis-idaea
1.03 6 0.03
0.81 6 0.03
47.83 6 0.30
45.15 6 0.16
46.64b 6 1.25
56.05a 6 2.31
Mosses
Aulacomnium turgidum
Hylocomium splendens
Sphagnum spp.
0.81 6 0.07
0.87 6 0.16
0.88 6 0.02
37.36 6 3.32
40.71 6 0.23
38.96 6 0.17
46.10b 6 1.01
46.61b 6 0.70
44.40b 6 0.71
Notes: Data are presented as means 6 SE. Different letters within the same foliar trait indicate that values are significantly
different at the P , 0.05 level from post-hoc tests after running one-way ANOVAs comparing each variable across species (n ¼ 6
individual replicates). Cell soluble refers to soluble cell contents (carbohydrates, lipids, pectin, starch, and soluble protein).
July 2014
SHRUB EXPANSION AND DECOMPOSITION
1869
FIG. 4. Leaf litter decay constants (k; means 6 SE) from (a) leaf litter collected at the moist acidic tundra species removal site
and decomposed for 5 yr in a common site, and (b) leaf and stem litter of different deciduous shrubs species collected across seven
different sites and decomposed for 3 yr in a common site. B. neoalaskana, the litter used in the common substrate experiment, was
also incubated in the common garden and is included in (b). Species are presented by functional group; deciduous shrubs (DESH),
evergreen shrubs (EVSH), graminoids (GRAM), and mosses. Full genus and species names are given in Table 1. Abbreviations
within the parentheses refer to the site where the litter was originally collected: Dalton Highway site 1 (D1), Dalton Highway site 2
(D2), Sagavanirktok River site 1 (S1), Sagavanirktok River site 2 (S2), low shrub site (L), medium shrub site (M), and high shrub
site (H). ANOVA results are from comparisons across species for samples collected at the species removal site. In panel (a), data
points that share a lowercase letter are not significantly different at the P , 0.05 level.
only weakly related to some of the litter quality indices
measured. The percentages of C and cell soluble
contents were positively correlated with decay rates (C;
r 2 ¼ 0.23, n ¼ 21, P ¼ 0.03, cell soluble contents; r 2 ¼
0.23, n ¼ 21, P ¼ 0.03, cellulose; r 2 ¼ 0.23, n ¼ 21, P ¼
0.04 [Appendix D: Fig. D1]). For stem litter, the
percentage of cellulose was the only litter quality index
that correlated positively with decay rates, explaining
55% of the variation in stem decay rate (Appendix D:
Fig. D2; r 2 ¼ 0.55, n ¼ 9, P ¼ 0.01).
TABLE 3. Extended.
Cell soluble (%)
Hemicellulose (%)
41.87b 6 2.61
35.79b 6 2.38
28.39a 6 1.25
30.82a 6 0.66
62.76a 6 0.75
64.98a 6 1.54
69.71a 6 0.11
63.29a 6 0.99
60.93a 6 0.75
Cellulose (%)
Lignin (%)
Lignin : N
24.47b 6 0.90
27.16a 6 0.46
4.81 6 0.93
4.30 6 0.37
3.60 6 0.64
4.91 6 0.43
10.15c,d 6 0.47
17.87b 6 1.12
9.49c,d 6 0.23
7.39e 6 0.47
9.82d,e 6 0.58
10.52c,d,e 6 0.09
19.44 6 0.31
6.78 6 0.57
8.35 6 0.36
11.72 6 0.29
3.54 6 0.22
4.83 6 0.33
8.14d 6 0.41
13.29c 6 0.85
12.14c,d 6 0.30
13.22c 6 0.27
16.08 6 0.36
11.19 6 1.45
15.66 6 0.38
14.09 6 2.26
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JENNIE DEMARCO ET AL.
Ecology, Vol. 95, No. 7
DISCUSSION
Microenvironment controls over litter decomposition
FIG. 5. (a) Aboveground net primary productivity (ANPP;
means 6 SE) across sites, and (b) mass loss (means 6 SE)
across sites with changes in litter quality and quantity alone and
with changes in litter quality, litter quantity, and soil
microenvironment. In each panel, data points that share a
letter are not significantly different at the P , 0.05 level, with
uppercase vs. lowercase letters in panel (b) indicating significant
differences across sites within the same variable.
Community-weighted mass loss
Community-weighted mass loss that took into account differences in litter quality and quantity across
sites was greatest in the high shrub site compared to the
medium and low shrub sites (site: F2,30 ¼ 3.5, P ¼ 0.05).
This was primarily driven by the higher overall ANPP
found at the high shrub sites compared to the medium
and low shrub sites (site: F2,30 ¼ 7.0, P , 0.01), because
the proportion that ANPP contributed to community
mass loss (mass loss/ANPP) was similar across sites (low
¼ 0.16, medium ¼ 0.15, and high ¼ 0.15). Communityweighted mass loss that took into account both
differences in litter quality and quantity and in the
environment was greater in the low and high shrub sites
than the medium shrub site (site: F2,30 ¼ 4.4, P ¼ 0.03;
Fig. 5) suggesting that different mechanisms may be
driving these differences at each of the sites.
Surprisingly, we were unable to detect any effect of
added snow on decomposition of the common litter
substrate after three years of incubation, despite a 28C
increase in soil temperature during the growing season
and a 48C increase during the winter. Walker et al.
(1999) also found no effect of deeper snow on litter
decomposition after two years of decomposing Betula
nana leaf litter under ambient and up to 3 m of added
snow in tussock tundra near Toolik Lake. In contrast, in
an alpine tundra community, Baptist et al. (2010) found
a trend for greater litter mass loss in late snowmelt sites,
presumably due to warmer soil temperatures in the
spring prior to snowmelt. Using the same species of litter
as our common litter experiment, Hobbie and Chapin
(1996) found differences in litter mass loss between
arctic tundra microsites whose summer soil temperatures differed by 48C, with greater mass loss occurring in
the warmer microsites. In addition, litter mass loss in lab
incubations that included warming treatments of either
28C or 68C above the ambient growing season temperatures showed an increase in mass loss with increased
temperature (Hobbie 1996, Jonasson et al. 2004). Two
out of the three studies had temperature differences that
were twice as high as ours, which may explain why they
found significant differences in litter mass loss with
change in temperature, while our study did not. At these
sites, temperature may not be the main control over
decomposition. Site-specific differences in soil moisture,
nutrient availability, litter quality and quantity, and the
decomposer community may play a larger role in
decomposition in these arctic and alpine communities.
When comparing decomposition of our common
substrate across our three sites, the common litter
decomposed faster in the low shrub site compared to
the medium and high shrub sites, even though ambient
soil temperatures at the medium and high shrubs sites
were actually warmer than at the low shrub site both
during the growing season and over the winter. Soil
temperature differences across these sites are of the same
magnitude as the differences in soil temperature we saw
when we added snow. Since we did not see strong
differences in decomposition when we added snow and
elevated soil temperatures, this suggests that other
factors such as moisture, soil nutrients, or the decomposer communities may be more important than small
(,48C) changes in temperature for driving decomposition at our sites. Soil moisture was not measured over
the 3-yr incubation period, but measurements in June of
2006 showed no difference in soil moisture across sites
(DeMarco et al. 2011) and could not explain the
differences we found in litter decomposition across the
sites.
Previous research from these sites has shown that
percentage of N of the top 10 cm of soil at the medium
and high shrub sites is twice as high as the percentage of
July 2014
SHRUB EXPANSION AND DECOMPOSITION
1871
FIG. 6. Soil N (means 6 SE) for each site vs. decay constant (k), initial C remaining, initial N remaining, and the proportion of
initial C to N remaining of the common substrate, B. neoalaskana, decomposed at each site over a 3-yr period.
N at the low shrub site at the same soil depth (DeMarco
et al. 2011) and is highly positively correlated with litter
initial N remaining (%) and initial C:N remaining (%),
but negatively correlated with decay constants (Fig. 6).
This suggests that soil nutrients may play an important
role in controlling litter decomposition and nutrient
release at our sites. Over our 3-yr incubation period,
litter decomposed at the low shrub site mineralized litter
N, while litter decomposed at the medium and high
shrub sites immobilized N. Thus, sites with greater soil
N have lower rates of decomposition and higher
retention of N on litter. In our study, sites that had
greater bulk soil N also had higher rates of net N
mineralization, suggesting greater N availability (DeMarco et al. 2011). Greater soil N availability has been
found to stimulate (Hobbie 1996, Aerts et al. 2006),
repress (Prescott 1995, Magill and Aber 1998, Aerts et
al. 2006), or have no effect (McClaugherty et al. 1985,
Prescott 1995, Hobbie 1996, Aerts et al. 2003, 2006) on
litter decomposition rates, and can lead to N immobilization in some systems (Gallardo and Merino 1992,
Magill and Aber 1998, Hobbie 2005, Aerts et al. 2006)
but see McClaugherty et al. (1985). These varying
responses of litter decomposition to external N may
have been attributed to interactions between the initial
quality of the litter and the availability of nitrogen in the
soil.
In a meta-analysis of 24 litter decomposition studies
in which external N was experimentally manipulated,
Knorr et al. (2005) found that external N availability
and litter quality interact to influence decay rates, with
N additions stimulating decomposition of high-quality
litters (,10% lignin content), while inhibiting decay of
low-quality (.20% lignin content) litters. The ‘‘microbial N mining’’ hypothesis suggests that this pattern
occurs because some microbes use labile C to decompose
recalcitrant organic matter in order to acquire N
(Fontaine and Barot 2005, Moorhead and Sinsabaugh
2006). Therefore, we would expect microbial N mining
to increase decomposition of low-quality litter when it is
incubated in soils with low soil N. Our highest
decomposition rates were seen at the low shrub site
where soil N (DeMarco et al. 2011) is low and soil
microbial activity is N-limited (Lavoie et al. 2011, Sistla
et al. 2012), suggesting that decomposition at this site
was driven by microbes mining for N. In contrast, in a
high soil N environment, nitrogen is readily available to
microbes, therefore it is not necessary for them to
mineralize litter to acquire N, which results in suppression of litter decomposition. Indeed there is evidence
that this can occur in laboratory incubations using leaf
litter from a range of plant functional types and multiple
soil types collected in southern Africa (Craine et al.
2007) and in field studies using leaf litter from the same
1872
JENNIE DEMARCO ET AL.
species but with varying litter quality and soil nutrient
availability (Talbot and Treseder 2012). The B. neoalaskana litter we used for our study was of low quality
with both a high C:N ratio (72 6 0.60; mean 6 SE) and
percentage of lignin (19 6 2) and was decomposed
across sites that varied in soil N availability (DeMarco
et al. 2011). Therefore the microbial N mining hypothesis may help explain why we saw a linear pattern of
decreased rates of decomposition and an increase in N
immobilization with increasing soil N at our sites.
The N immobilization we found at our high shrub site
may also be explained by interactions with soil nutrient
availability and the high lignin content of our litter.
High nutrient content in soils can suppress the
production of fungal ligninase (Carreiro et al. 2000,
Sinsabaugh et al. 2002), which is induced by low N
availability (Keyser et al. 1978), resulting in low rates of
decomposition. Fungal communities between tussock
tundra and shrub tundra soils sampled near Toolik Lake
differ at the phyla and subphyla levels; however, we do
not know whether the species responsible for breaking
down lignin or the production of ligninase differs
between these two plant communities (Wallenstein et
al. 2007). These tussock tundra soils are dominated by
slow-growing microbes that have high affinities for C
substrates of low quality and quantity. In contrast,
shrub tundra soils are dominated by microbes that have
high growth rates with high nutritional requirements for
C substrates of higher quality and quantity (Fierer et al.
2007, Wallenstein et al. 2007). It is possible that
microbes at the low shrub site are better at decomposing
litter that is of low quality than the microbes at the high
shrub site, and that microbes at the high shrub site
immobilize more N because they have a higher nutrient
demand when mineralizing C. Low-quality litters can
also contain high tannin contents, which can bind to N
and become incorporated in the lignin fraction, decreasing decomposition and increasing immobilization of N
(Gallardo and Merino 1992, Aerts et al. 2003). Although
we did not measure tannins in our litter, others have
found that Betula spp. leaves can have more polyphenols
than Salix spp. and Populus spp. leaves (Palo 1984). In
addition, litter of B. papyrifera grown under elevated
CO2 increased tannin content by as much as 81%. When
decomposed in a common garden, this high tannin
content litter had lower rates of decomposition and
higher N immobilization compared to litter from
ambient CO2 conditions, which had lower tannin levels
(Parsons et al. 2004).
Results from our common substrate experiment
suggest that, at least on a short time scale, any effect
shrubs have on soil microclimate via their ability to trap
snow will have little influence over nutrient turnover
through the decomposition of aboveground litter. How
shrubs influence decomposition of belowground litter
(roots and rhizomes) remains uncertain. In contrast,
changes in soil nutrient availability and/or changes in
soil microbial community associated with an increase in
Ecology, Vol. 95, No. 7
deciduous shrub abundance will likely decrease the rate
at which litter decomposes, assuming that the native
litter responds similarly to the common substrate we
used in our experiment. The litter of B. neoalaskana had
a higher C:N ratio than the deciduous shrub litter native
to these sites but similar lignin content; thus, we assume
that when native litter is decomposed at the high shrub
site it would respond in the same way as our common
substrate.
Litter quality controls over litter decomposition
Members of a plant functional group (i.e., deciduous,
evergreen, graminoid, etc.) are often similar in their litter
chemistry and decomposition rates (Hobbie 1996,
Hector et al. 2000, Hobbie and Gough 2004). Comparisons of litter decomposition among species and
functional groups in this experiment suggest that
decomposition rates of arctic plants cannot be generalized using functional group designations, because
species within the same functional group did not always
follow the same pattern of decomposition. For example,
the evergreen shrub, Ledum decumbens, and the graminoid, Eriophorum vaginatum, had decomposition rates
that were similar to those of the deciduous shrubs, B.
nana and Vaccinium uliginosum. Their decomposition
rates were higher than other species within their
functional groups; Ledum decumbens had higher rates
than the evergreen shrub Vaccinium vitis-idaea, and
Eriophorum vaginatum had higher rates than the
graminoid Carex bigelowii. This is in contrast with
other decomposition studies within this region (Hobbie
1996, Hobbie and Gough 2004), perhaps because our
study included more species of deciduous and evergreen
shrubs than were used in previous studies. Decomposition rates varied significantly among species. Of the 10
vascular plant species we decomposed, Rubus chamaemorus had the highest quality litter and the fastest rate
of decomposition, losing about 70% of its mass over a 3yr period. Aerts et al. (2006) also found that R.
chamaemorus decomposed more quickly than three
other subarctic bog species. In contrast, mosses had
the lowest decomposition rates, only losing about 30%
of their mass over the same period. Differences among
the rest of the species were relatively small. A change in
species composition in the Arctic could lead to
alterations in community decomposition rates and
nutrient turnover if the change includes an increase in
the relative abundance of R. chamaemorus and a
decrease in mosses as seen in fertilized tussock tundra
in Alaska where R. chamaemorus dominates the
understory and moss cover is reduced (Chapin et al.
1995).
Litter quality and quantity vs. microclimate controls
over litter decomposition
Based on our results, changes in litter quality and
quantity via shrub expansion will increase total community-weighted mass loss. This appears to be primarily
July 2014
SHRUB EXPANSION AND DECOMPOSITION
driven by changes in litter quantity, rather than litter
quality and microenvironment. For example, the low
and medium shrub sites had similar ANPP rates and
similar mass loss rates, even when differences in litter
quality among the sites were taken into account. This
suggests that either the quality of litter among the two
sites is not different enough to cause differences in
community mass loss, or that litter quality is not the
main driver of decomposition at these two sites. Of the
three sites, the high shrub site had the highest ANPP and
the highest mass loss, suggesting that the high quantity
of litter produced at the high shrub site is driving the
large mass loss from this site. The magnitude of
difference in mass loss among the sites, however, is
slightly minimized if changes in the soil microenvironment are taken into account, with a trend for an increase
in mass loss at the low shrub site and a decrease in mass
loss with increasing shrub abundance. This suggests that
community-weighted mass loss is more sensitive to
shrub-induced changes in the amount of litter inputs
than to shrub-induced changes in the soil microenvironment.
Carbon loss of our litter closely followed that of mass
loss (Fig. 2). Initial percentage of C of our litter was
between 40% and 45%. Therefore, we can assume that
community C loss would follow a similar pattern as
mass loss, with the absolute values of C loss being about
half those of mass loss. Based on these assumptions and
our data, shrub expansion will lead to about a 55%
increase in C loss from litter. However, we do not know
the exact fate of the C lost from the litter, which can be
released back to the atmosphere as carbon dioxide,
leached into the soil as dissolved organic C, or be
incorporated as soil organic matter and microbial
biomass. Previous research at these sites show that soil
C pools in the top 10 cm at the medium and high shrub
sites have 40–60% more C stored in them compared to
the soil C pool at the low shrub site (DeMarco et al.
2011). This suggests that a larger portion of the C lost
from litter is most likely remaining in the soil at the
medium and high shrub sites compared to the low shrub
site.
We assumed that litter from all species within shrub
communities will respond to the environment in the
same ways as the common litter we incubated. However,
it is possible that differences in litter quality may interact
with site environment at the high nutrient site, resulting
in a different pattern than what we saw with our
common litter. Litter quality and site interactions are
not always reported, as many decomposition studies do
not conduct a full factorial experiment where all litter
types are incubated in all environments. However,
Hobbie and Gough (2004) did show that B. nana stem
decomposition followed the opposite pattern when
incubated in moist acidic and moist nonacidic tundra
compared to the other species decomposed at the same
sites.
1873
CONCLUSIONS
Our study suggests that on a short time scale, the
changes in soil temperature and moisture associated
with additional snow trapping by shrubs are unlikely to
influence litter nutrient turnover enough to drive the
positive snow–shrub feedbacks proposed by Sturm et al.
(2001). The mechanisms driving shrub expansion are
more likely to do with shrub–litter feedbacks, where the
higher growth rates and N uptake by shrubs allows them
to produce more leaves, resulting in a larger litter N pool
and faster internal cycling of nutrients (DeMarco et al.
2011). Retention of N in litter during the early stages of
decomposition in shrub sites may be beneficial for soil
organic matter decomposition and could help explain
why we see more soil N and greater N mineralization at
the medium and high shrub sites.
ACKNOWLEDGMENTS
We thank Charmagne Wasykowski, Grace Crummer, Anne
Baker, Rafael Mendoza, Yi Wei Chang, Hanna Lee, Caitlin
Hicks-Pries, and the many undergraduates at the University of
Florida for their assistance in the field and in the lab. This
research was supported by NSF grants DEB-0516041, DEB0516509, OPP-6737545, and the Arctic LTER (DEB-0423385).
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SUPPLEMENTAL MATERIAL
Appendix A
Aboveground net primary productivity for each species within each shrub community and their corresponding k values
(Ecological Archives E095-164-A1).
Appendix B
Initial mass, carbon, and nitrogen remaining, and decay constants for the common substrate, Betula neoalaskana (Ecological
Archives E095-164-A2).
Appendix C
Initial litter quality of leaves and stems of Betula nana and Salix pulchra collected across the shrub gradient sites (Ecological
Archives E095-164-A3).
Appendix D
Regression analysis of initial litter quality and litter decay constants (Ecological Archives E095-164-A4).