Introduction
Soils with active biological soil crust (biocrust) communities are essential
components of dryland ecosystems worldwide and are also one of the most
sensitive components of drylands to climate change (Ferrenberg et al., 2017;
Reed et al., 2016). Given the vast and growing global extent of dryland
regions (Safriel et al., 2005; Prăvălie, 2016), the response of
biocrusts to major global change phenomena, such as climate change, may be an
important aspect of the overall response of Earth's ecosystems. In
particular, due to the potential for dryland feedbacks to future climate
(Poulter et al., 2014; Ahlström et al., 2015; Rutherford et al., 2017), a
key parameter to consider as dryland ecosystems warm is carbon (C) balance,
specifically carbon exchange of biocrusted soils. Dryland soils are
characterized by low soil organic matter that is negatively correlated with
aridity across many drylands (Delgado-Baquerizo et al., 2013) and there is an
association between C loss and the phenomenon of desertification (Lal, 2004).
Drylands can also show large year-to-year variation in C fluxes that are
relevant for explaining global-scale fluxes (Ahlström et al., 2015;
Poulter et al., 2014; Biederman et al., 2017). At the ecosystem scale,
biocrusted soils within drylands are often substantial contributors to both C
uptake (Elbert et al., 2012) and ecosystem respiration (Castillo-Monroy et
al., 2011). At the organism scale, the viability of biocrust is linked to
their ability to maintain a positive C balance among hydration–desiccation
cycles (Grote et al., 2010; Coe et al., 2012; Oliver et al., 2005). Despite
the importance of C balance to understanding biocrust function and dryland
ecosystem feedbacks to global change, few studies have addressed how biocrust
soil CO2 fluxes will respond to changing temperature and
precipitation.
Carbon balance in biocrusted soils includes not only the activities of the
biocrusts themselves, but also the activities of subsurface vascular plant
roots and soil heterotrophic microbes. Considering biocrusted soils together
with the function of adjacent vascular plants is important given that there
is increasing evidence for biotic connections, possibly mediated by fungi,
between these functional groups (Green et al., 2008) and for linkages in
plant–soil C cycle responses to warming. For example, at another site on the
Colorado Plateau, measurements of plant photosynthesis, coupled with spot
measurements of soil respiration under plant canopies, showed plant
photosynthetic rates were tightly coupled to soil respiration rates, with
both showing reduced fluxes in response to warming during the spring when
plants are most active (Wertin et al., 2017). While these patterns could be
the result of independent climate controls, such as temperature and moisture,
on each individual flux, vascular plant C allocation to roots and
heterotrophs belowground or biotic connections between biocrust organisms and
vascular plants could also help explain the coupling between above- and
belowground CO2 fluxes.
In addition to affecting soil C balance through direct physiological means,
warming has been shown to have substantial effects on biocrust species
composition, including macroscopic components such as moss and lichens
(Ferrenberg et al., 2015; Escolar et al., 2012; Maestre et al., 2015) and
microbial communities (Steven et al., 2015; Johnson et al., 2012). Climate
models predict rapidly rising temperatures for already hot and
moisture-limited dryland regions, including the site of our study in the
southwestern United States (IPCC, 2013; Jardine et al., 2013). Forecasts
of future precipitation patterns are less certain, but overall drier
conditions with changes in precipitation event size and frequency are likely
(Seager et al., 2007). Climate models predict increases in dryland annual
average temperature of up to 4 ∘C by the end of the 21st century, as
well as significant alterations to the amount and timing of rainfall
(Christensen et al., 2007). For example, the Intergovernmental Panel on
Climate Change (IPCC) A1B scenario suggests a decrease in precipitation
amount of 5–10 % for the southwestern United States, as well as significant changes
to the timing and magnitude of precipitation (D'Odorico and Bhattachan,
2012). Across many ecosystems, including drylands, both plant C uptake and
soil respiration show an optimum, such that rates are positively correlated
with increased temperatures and moisture (Wu et al., 2011) until a point at
which high temperatures (often accompanied with drying) begin suppressing
both photosynthesis (e.g., Wertin et al., 2015) and soil respiration (Tucker
and Reed, 2016). Drought also tends to reduce vascular plant production and
respiration, with greater sensitivity in drier areas (Knapp et al., 2015). In
soils overlain by biocrusts (hereafter, biocrusted soils) specifically,
temperature and moisture are key physiological parameters for C flux (Grote
et al., 2010; Darrouzet-Nardi et al., 2015) and, although few, the warming
experiments that do exist suggest that biocrusted soils will have higher net
CO2 efflux with a warming climate (Darrouzet-Nardi et al., 2015;
Maestre et al., 2013). There is evidence for a limit to this association
though, with very high temperatures leading to reduced biotic activity,
including microbial respiration, in biocrusted soils (Tucker and Reed, 2016).
MAT is the mean annual temperature. Values are shown for the nearby
Moab site (see Fig. S2 for long-term record) as well as for the instruments
at our study site. Values in parentheses indicate the number of days of
missing data for the given year. MAP is the mean annual precipitation and
spring precipitation totals were determined by a rain gauge at the study
site. Detailed timing of temperature and precipitation over the study period
are shown in Fig. S1. Supplemental water was only added to the watering and
combined treatments and was not added on days when natural precipitation
occurred. Spring rainfall is from day of year 80–173 and is the time of peak
plant growth.
Year
Moab
Moab
Study site
Study site
Spring
Supplemental
First
Last
Number of
MAT
MAP
MAT
MAP
precipitation
water
watering
watering
watering
(∘C)
(mm)
(∘C)
(mm)
(mm)
(mm)
date
date
days
2006
22.6 (2)
208 (0)
21.4 (0)
294 (0)
22
48
31 May
20 Sep
40
2007
22.9 (8)
191 (4)
22.1 (0)
223 (0)
68
42
14 Jun
20 Sep
36
2008
21.8 (4)
138 (0)
22.6 (0)
200 (0)
62
44.4
17 Jun
23 Sep
43
2009
21.9 (1)
126 (0)
20.8 (1)
189 (0)
57
27.8
10 Jun
4 Sep
32
2010
21.4 (0)
204 (0)
20.0 (13)
286 (13)
51
48
9 Jun
29 Sep
40
2011
21.7 (0)
161 (0)
20.0 (1)
199 (0)
71
42
13 Jun
19 Sep
36
2012
23.6 (1)
92 (1)
22.1 (85)
122 (84)
9
54
4 Jun
5 Oct
45
2013
20.7 (2)
183 (2)
19.3 (36)
253 (32)
43
0
31 May
20 Sep
0
2014
22.8 (0)
208 (0)
21.5 (1)
304 (0)
73
0
14 Jun
20 Sep
0
To improve our understanding of dryland C flux responses to global change, we
used a warming by watering manipulation experiment on the Colorado Plateau
established in 2005. When the study began, we explored the hypothesis that
warming would increase net losses of CO2 from soils covered with late
successional biocrusts (∼50 % moss, ∼30 % lichen cover)
via detrimental impacts on biocrust physiology caused by warming. At the same
time we wanted to explore how altered precipitation could directly affect
biocrust soil CO2 exchange and/or interact with the effects of
increased temperatures. These early results supported the basic hypothesis
concerning the warming-only treatment, showing that warming led to increased
CO2 loss after 1–2 years, with the largest differences during
periods in which soils were wet enough to support substantial biocrust
photosynthesis (Darrouzet-Nardi et al., 2015). Crucially, we also found that
the increased frequency of small frequent precipitation events negatively
affected biocrusts: the treatment caused the death of a major biocrust
component, the moss Syntrichia caninervis (Coe et al., 2012; Reed et
al., 2012; Zelikova et al., 2012). This finding represented a substantial
alteration to the system and led to a second phase of the experiment. In this
phase, we ceased the watering treatment that had caused moss death and
increased the warming treatment from 2 to 4∘ to see if greater
warming would negatively impact biocrusts. We found that the greater warming
did in fact reduce moss and lichen cover as well, though not as rapidly as
the watering treatment (Ferrenberg et al., 2015). Here we report the C
balance response to these multiple phases of the experiment. Our main goals
were to (1) determine if the increased net soil CO2 loss observed
after a year of warming was maintained after 8 years and (2) to assess how
the altered precipitation patterns affected net soil CO2 exchange
during the early phase when mosses were dying and, then later, after mosses
were lost and the increased watering had ceased.
Materials and methods
Site description
The study was located in a semiarid ecosystem on the Colorado Plateau
(36.675∘ N, -109.416∘ W; elevation = 1310 m; mean annual
temperature = 13 ∘C, mean annual precipitation = 269 mm;
WRCC 2014) that supports multiple species of grasses and shrubs. Soils are
Rizno series Aridisols and the dominant plants include Achnatherum hymenoides, Pleuraphis jamesii, Atriplex confertifolia, and Bromus tectorum. Biocrust communities are dominated by the cyanobacterium
Microcoleus vaginatus the moss Syntrichia caninervis, and
the cyanolichens Collema tenax and Collema coccophorum. The
site is on a moderate hillslope (∼10 %) surrounded by steep gullies
that make it hard to access for livestock, which may explain its relict
biocrust and plant composition that includes late successional crusts with
well-developed communities of native grasses and shrubs, similar to sites
found in Canyonlands National Park (Belnap and Phillips, 2001). Rainfall
during the study period was distributed around the mean (Table 1, Fig. S1 in
the Supplement), with several slightly above average years including the
first and last year of the experiment (2006: 294 mm; 2014: 304 mm)
and with 1 year with substantial drought (2012: 122 mm). Rainfall and temperatures went
up and down across years, with no notable directional shift over the 9-year
course of the study. Long-term records from a nearby weather station in Moab,
UT, show that mean annual temperatures have been increasing (21.3 ∘C
for 1900–1924 vs. 22.9 ∘C for 1991–2016, a difference of
1.5 ∘C). Precipitation trends since 1925 do not show a clear trend
(Fig. S2).
Warming and watering treatments
The experiment contained 20 plots with five replicates (n=5) for each of
four treatments: control, warmed, watered, and
combined (warmed + watered). Plots were 2×2.5 m in
size and grouped into five blocks determined by spatial location on the
hillslope. Each plot contained one automated CO2 chamber (described
below). The warming treatment began in October 2005 in plots fitted with
800 W infrared radiant (IR) heat lamps (Kalglo model MRM-2408) mounted at a
height of 1.3 m. Control plots had dummy lamps that do not provide heat. The
heating treatment was regulated by altering the voltage supplied to each
lamp. While some drying of soil moisture from the lamps may have occurred, we
saw little evidence for this phenomenon in soil moisture values, with drying
after precipitation events occurring at similar rates in all treatments
(Fig. S3). A previously published analysis also reported no easily detectable
moisture effects from the infrared lamps in either this experiment or a
similar co-located experiment despite soil moisture probes at 2, 5, and
10 cm throughout all plots (Wertin et al., 2015). However, we cannot rule
out very shallow surface moisture effects, which could be important (Tucker
et al., 2017).
The target temperature increase was ambient soil temperature +2 ∘C
from 2005 to 2008, at which point a second lamp was added to each plot and the
warming treatment was increased to +4 ∘C, where it remained through
the end of the automated chamber sampling in September 2014. The treatment
temperatures were increased from 2 to 4 ∘C above ambient in order to
better match changing predictions of future temperature by 2100 (Christensen
et al., 2007). To simulate predictions of increased frequency of small
precipitation events (Weltzin et al., 2003; Christensen et al., 2007), water
was added in 1.2 mm events manually with backpack sprayers and was applied
40 times from 31 May to 20 September 2006 and 36 times from
14 June to 20 September 2007, with an average time between watering of 2.8 days
(∼4 × natural frequency; Table 1). This watering treatment
continued through 2012 (Table 1). The amount of water varied by year because
watering did not occur on days when natural rainfall occurred. Watering was
stopped after 2012 because the late successional biocrust community had been
eliminated after the first year and was showing no further change through
time (Reed et al., 2012; Ferrenberg et al., 2015).
Net soil exchange measurements with automated chambers
Carbon dioxide fluxes were assessed with automated CO2 flux chambers,
described in detail in Darrouzet-Nardi et al. (2015). The chambers were
placed within the soil, open at the bottom, and have clear lids at the top
that are closed once per hour for 3 min to assess net CO2 flux. The
chambers allow in sunlight and hence allow photosynthesis by biocrust
organisms. Fluxes of CO2 during that time are calculated as the rate
of change in CO2 concentrations during the 3 min period. During that
3 min period, CO2 was recorded every 2 s and averaged every 10 s.
Aberrant points were down-weighted with a smoothing function (“supsmu”
implemented in MATLAB; Friedman, 1984), allowing a robust calculation of
slope for a given 3 min interval (Bowling et al., 2011). The chambers were
30 cm tall × 38 cm inner diameter, covering a soil surface area of
0.11 m2. Chambers were installed to a depth of 27 cm in the soil,
leaving ∼3 cm of the chamber protruding above the soil surface. The
chambers were placed in plot locations containing biocrusts but no vascular
plants. Values from these chambers were reported as net soil exchange (NSE)
of CO2. The concept of NSE is defined in Darrouzet-Nardi et
al. (2015) to include biocrust photosynthesis as the sole form of CO2
uptake (i.e., because the chambers do not include vascular plants) along with
CO2 losses via respiration from biocrusts, other soil microbes, plant
roots, and any abiotic soil sources. While it would have been ideal to
operate the chambers year round for the entire course of the experiment, it
was beyond the operational capacity of the project to do so and there are
times when the systems were not operational. The chambers have more frequent
malfunctions during the winter due to weather conditions, so those months are
least represented. There were intermittent automated chamber measurements in
2012, the last year of watering, crossed with the higher warming level,
providing enough data for analyses of daily patterns, though not enough to
assess seasonal total rates.
Biocrust community composition of the autochambers was measured at the
initiation of the experiment in 2005 and again in 2017. Assessment of the
biocrust community was performed using a frame that covered the autochamber
area in which the cover of 31 individual 25.8 cm2 squares was
estimated for all biocrust species.
The total cover of each species was
summed from the individual quadrats and the quadrats covered 800 cm2 of
the chambers' 1100 cm2 area.
Cover (%) of major biocrust constituents inside of the automated
CO2 flux chambers representative of the early and later periods of
the study.
(a) The 24 h average net soil exchange (NSE) of
CO2 through all treatments and years. Dates of supplemental watering
applications are shown as vertical blue lines. Ribbons indicate ±1 SE.
Precipitation is shown above each year's data, with annual totals shown on
the left and the size of several of the largest events noted for scale. Means
for each treatment are shown with different colors representing different
treatments (control = black, warmed = red, altered monsoonal
precipitation [watered] = blue, warmed × watered
[combined] = purple). Positive NSE rates depict respiratory losses that
were greater than CO2 uptake and negative NSE rates depict C fixation
rates that outpaced respiratory losses. (b) Differences between
treatments and control (td) are shown as solid lines ±95 % CI calculated for each daily average shown with shading. Values
were calculated by subtracting the control rates from the treatment
(red = warmed - control; blue = altered monsoonal precipitation
[watered] - control; purple = warmed × watered
[combined] - control).
Imputation and statistical analysis
Hourly data from the automated chambers were collected from
1 January 2006–20 September 2007, 19 February–17 November 2013, and
14 February–17 November 2014, for a total of 28 058 time points for each of
the 20 chambers. Of these time points, 29 % of the data were missing,
primarily due to technical issues with the chambers. To allow calculations of
cumulative NSE, data were imputed following the same procedure as in our
previous work (Darrouzet-Nardi et al., 2015). Data were assembled into a data
frame containing columns for (i) each of the 20 chambers; (ii) environmental
data including soil and air temperature, soil moisture, 24 h rainfall
totals, and photosynthetically active radiation (PAR); and (iii) 6 days of
time-shifted fluxes (before and after each measurement; i.e., -72, -48,
-24, +24, +48, +72 h) for one chamber from each treatment, soil
temperature, and soil moisture. Lagged values were added due to their ability
to greatly improve prediction of missing time points, particularly for short
time intervals such as those caused by, for example, several hours of power
outage at the site. One data frame was created for each of the three
continuous recording periods – 2006–2007, 2013, and 2014 – and each was imputed
separately. Imputation was performed using the missForest algorithm, which
iteratively fills missing data in all columns of a data frame using
predictions based on random forest models (Stekhoven and Buhlmann, 2012;
Breiman, 2001).
After imputing the hourly values, cumulative fluxes were calculated by
summing NSE over 7-month periods (19 February–19 September) for each
year (2006, 2007, 2013, and 2014). This 7-month period was selected due
to availability of data in all four analysis years. The total number of
cumulative fluxes evaluated was 80
(4 years × 4 treatments × 5 replicates). We also made
separate cumulative estimates of time periods in which we observed active
photosynthesis, defining these periods as days during which the NSE values
were -0.2 µmol CO2 m-2 s-1 or lower, with
more negative numbers showing higher net photosynthesis. These periods
typically correspond to times with sufficient precipitation to activate
biocrusts. The effect of the warmed, watered, and
combined treatments on cumulative NSE values was evaluated by
calculating the size of the differences between each treatment and the
control (Nakagawa and Cuthill, 2007; Cumming, 2013). Treatment differences,
which we notate as td, were calculated as treatment - control
(paired by block) with 95 % confidence intervals estimated using mixed
effects linear models for each year with treatment as a fixed effect and
block as random effect (Pinheiro and Bates, 2000). Analyses were facilitated
by a custom-made R package “treateffect”, available at
https://github.com/anthonydn/treateffect (2bb5ed2, last access: 7 May 2018). The data used for these analyses are available
at 10.6084/m9.figshare.6347741.v2. Finally, to evaluate differences
over time, differences between 2006 data for each treatment and each
subsequent year were calculated, also using mixed effects models.
Results
Biocrust cover within the soil collars used by the automated chambers was
relatively similar in all treatments at the beginning of the experiment, with
an average of 49 % moss and 31 % lichen in each treatment (Fig. 1).
Between 2005 and 2017, these percentages fell in all treatments including the
controls, eventually being replaced primarily by lightly pigmented
cyanobacterial crusts, probably Microcoleus vaginatus (Gundlapally
and Garcia-Pichel, 2006). Lichen went to <3 % in all treatments.
Mosses were more variable, remaining at 25 % in controls, but falling to
7 % in warmed plots and to 0 % in both watering plots. Cyanobacteria
cover started at 0 % in all chambers and rose to 50–90 %.
Effect sizes of our treatments are shown as mean differences in NSE
between treatments and controls with 95 % confidence intervals
(td). Values were calculated as the control plot rate subtracted
from the rate in the treatment plot, with positive values indicating higher
NSE values in the treatment plot relative to the control and vice versa.
Analyses correspond to the NSE data shown in Fig. 4. Note that all underlying
fluxes are positive (source to atmosphere), but here the differences
between treatments are shown.
Seven-month
Active photosynthesis
periods
periods
Year
Comparison
td (g C m-2)
td (g C m-2)
2006
warmed - control
5.1 [-9.7, 19.9]
4.1 [-0.1, 8.2]
2006
watered - control
14.6 [-0.2, 29.4]
5 [0.8, 9.1]
2006
combined - control
9.8 [-5.1, 24.6]
7.6 [3.5, 11.8]
2007
warmed - control
6.1 [-6.7, 18.7]
2 [0.6, 3.5]
2007
watered - control
10.9 [-1.8, 23.6]
1.5 [0, 2.9]
2007
combined - control
8.33 [-4.4, 21.0]
2.6 [1.2, 4.1]
2013
warmed - control
-10.7 [-27.7, 6.2]
1.3 [-0.5, 3.1]
2013
watered - control
-15.3 [-32.2, 1.6]
-0.1 [-1.8, 1.7]
2013
combined - control
-11.8 [-28.7, 5.2]
0.9 [-0.9, 2.7]
2014
warmed - control
-1.2 [-30.6, 28.1]
2.9 [-1.1, 7]
2014
watered - control
-4.0 [-33.3, 25.3]
0.4 [-3.7, 4.4]
2014
combined - control
-6.2 [-35.5, 23.1]
1.6 [-2.4, 5.6]
Seasonal time courses of NSE showed similar patterns among years and
treatments, with peaks in NSE in the spring associated with peak vascular
plant activity and peaks in both negative and positive NSE associated with
rain events (Fig. 2a). In the early time period (1–2 years after treatments
began), the supplemental 1.2 mm watering treatment caused large “puffs” of
CO2 when water was added. By the final year of watering (2012), the
size of these puffs was substantially smaller and after watering ceased
(2014); they did not occur even with natural rainfall events (Fig. 3).
In the early time period (2006–2007), interannual comparisons of cumulative
19 February–19 September (7-month) CO2 fluxes were consistent
with the hypothesized trend of the warming and watering treatments increasing
CO2 flux to the atmosphere. In the early time period, shortly after
the establishment of the treatments, we observed higher NSE (greater movement
of CO2 from soil to the atmosphere) in both watered and combined
treatment plots, with less evidence of difference in the warming-only
treatment (Fig. 4a; Table 2). Fluxes were similar between 2006 and 2007
(Table S1 in the Supplement).
Interannual comparison of “puffs” of CO2 from single
automated flux chambers (watering treatment, block 2 in blue and comparable
control chambers in gray) observed in response to midsummer experimental
watering treatments. Time resolution is hourly. Plots were experimentally
watered from 2005 to 2012, with no watering in the final panel (2014). Timing
of the watering treatments is shown by the vertical dotted lines. The puffs
shown here are CO2 fluxes at or above ∼1 µmol CO2 m2 s-1 and these occurred in
response to active watering treatments.
In the later time period (2013–2014), the treatments showed varying results.
In 2013, after the watering treatment had ceased, we observed a reversal of
the treatment trend from the early period, with lower CO2 efflux from
soils in all three treatments (Fig. 4a; Table 2). This trend was particularly
visible in the months of May and June (Fig. 2a, b). However, in the following
year, 2014, a wet year with high spring rainfall (Table 1, Fig. 2a), all
plots showed the highest CO2 efflux observed in the experiment (e.g.,
36.2 [21.7, 52.9] µmol m-2 s-1 higher compared to 2006
in control plots; Table S1). While no obvious treatment effects were
observed, treatment effect sizes were relatively poorly constrained due to
the higher variation that year (Table 2).
(a) The 7-month cumulative CO2 fluxes during 4
measurement years – 2006, 2007, 2013, and 2014 – for the period of
19 February–18 September, a period chosen due to availability of data in all
measurement years. (b) Cumulative CO2 flux during periods
with active photosynthesis (defined as days during which NSE was <-0.2 µmol CO2 m-2 s-1 or lower, largely
corresponding with wet periods). Though selection was made on this daily
minimum, numbers are positive because 24 h totals during these periods were
still largely net sources of CO2 to the atmosphere despite active
photosynthesis during peak hours. Dots indicate values from individual
automated chambers and horizontal and vertical bars indicate mean ±SE.
For effect sizes associated with each treatment, see Table 2.
Interannual comparisons of cumulative CO2 fluxes during periods of
active photosynthesis showed higher photosynthesis in all treatments during
the early measurement period (e.g., 2006 warmed td=4.1 [-0.1,
8.2]; Fig. 4b; Table 2). In the later period (8–9 years after treatments
began), subsequent to the cessation of watering, warmed plots still showed
elevated CO2 losses during periods of active photosynthesis but this
difference was smaller than in the earlier measurements (e.g., 2013 warmed
td=1.3 [-0.5, 3.1]; Fig. 4b; Table 2). In contrast, watered
plots that were not warmed were similar to control plots.
In examining the daily cycles in the hourly data, further detail on the
nature of the treatment effects was observed. After 1 year, watered
treatments in which mosses had died showed strong reductions in CO2
uptake capacity during wet-up events, but warmed treatments still showed a
similar maximum uptake capacity relative to controls (e.g., minimum NSE on
15 October 2006 control = -0.93±0.19 µmol m2 s-1; warmed = -0.89±0.11,
watered = -0.35±0.06, combined = -0.2±0.08; Fig. 5a).
However, after 8 years of treatment, clear differences were present in the
CO2 flux dynamics in response to natural rainfall events (Fig. 5b).
Biocrusted soils in control plots still exhibited substantial net uptake of
CO2 (e.g., minimum NSE on 14 August, control = -0.68±0.12 µmol m2 s-1), whereas the other treatments showed
less uptake relative to the control, with a similar trend visible on
23 August.
Examples of hourly CO2 flux patterns during rain
events (a) early in the experiment and (b) in the final
season of measurement. Solid lines are the mean and ribbons indicate ±1 SE. See Fig. 1a for rainfall patterns at these times.
Discussion
Early period: 2 ∘C warming × watering
(2006–2007)
The increase in CO2 effluxes in the watered treatments during the
early period (Fig. 4, Table 2) were likely driven by both the loss of
photosynthetic biocrust organisms during that time (Reed et al., 2012) and
increased soil respiration from soil heterotrophs. Moss death may
have contributed to net soil C loss via (i) eliminating CO2 uptake
from this important biocrust CO2 fixer (Reed et al., 2012; Coe et
al., 2012) and (ii) decomposition of dead mosses. Elevated soil respiration
with warming and watering is broadly consistent with the results of similar
experiments across many ecosystems (Wu et al., 2011; Rustad et al., 2001),
dryland sites specifically (Nielsen and Ball, 2015; López-Ballesteros et
al., 2016; Patrick et al., 2007; Thomey et al., 2011), and previously
documented effects in biocrusted soils at this site and others
(Darrouzet-Nardi et al., 2015; Maestre et al., 2013; Escolar et al., 2015).
In the warmed treatment, elevated NSE was not as evident in 2006 as in the
watered and combined treatments, and this is consistent with the biocrust
community changes. While moss died off quickly in the watered plots, mosses
in the warmed plots took longer to show negative effects (Ferrenberg et al.,
2017). Indeed, increased CO2 efflux with warming was clearer in the
following year (2007) and moss cover was substantially reduced by 2010
(Ferrenberg et al., 2015). Such rapid species composition changes have been
repeatedly implicated as drivers of system change in drylands, even with
seemingly subtle changes in climate (Wu et al., 2012; Collins et al., 2010).
Late period: 9-year warming (2–4 ∘C) × legacy
watering (2013–2014)
During the later period (2013) when warming had been increased to
+4 ∘C (in 2009) and watering had ceased (effectively making the
treatments control, +4 ∘C, legacy watering, and
+4 ∘C × legacy watering), several differences in
treatment effects emerged in comparison to the early measurement period
(2006–2007). First, the trend in the 2013 7-month cumulative CO2
fluxes (Fig. 4, Table 2) was reversed from those of the early measurement
period (2006–2007), with the control plots having the highest NSE and all
other treatments showing lower CO2 efflux. The reversal of the NSE
trend in the +4 and +4 ∘C × legacy watering treatments
is likely influenced by changes in biocrust community composition, with
mosses largely eliminated in relation to the control plots where about half
of the mosses were retained (Fig. 1). By 2013, lower NSE in warmed and
watered plots may have been linked to the completion of moss and lichen
decline and thus cessation of fluxes from sources such as decomposition or
exudation. Reductions in biocrust cover were also observed in the control
plots perhaps due to the longer-term effects of infrastructure, human
variation in community assessment, or natural variation in community
composition (Belnap et al., 2006), and such changes could help explain the
higher NSE in controls in 2013. Another possibility is that the reduced
vascular plant photosynthesis observed for multiple plant species with
warming in this area (Wertin et al., 2015, 2017) reduced plant allocation of
C belowground. This trend could reduce root C efflux and heterotrophic
breakdown of root exudate C, leading to the observed lower NSE values. A
number of warming experiments in more mesic systems that do not have
photosynthetic soils have shown an initial warming-induced increase in soil
CO2 respiratory loss followed by subsequent declines in warmed plots;
in these situations, reduced soil C availability for heterotrophic
respiration and changes to heterotroph C use efficiency are often suggested
to play a role (Bradford et al., 2008; Bradford, 2013; Tucker et al., 2013).
Such effects would also be consistent with drying from the infrared heat
lamps, a mechanism that was supported in a Wyoming grassland experiment
(Pendall et al., 2013). Our soil moisture data showed little evidence of such
drying effects (Fig. S3). However, with a minimum moisture probe depth of
2 cm, we may have missed moisture effects relevant only to the top several
millimeters of soil, an area of current active investigation at the site:
more recent results suggest that surface moisture (0–2 mm) can be a potent
predictor of soil C fluxes on these biocrusted soils (Tucker et al., 2017).
The reduction in CO2 efflux with warming was also seen in a nearby
set of plots in 2011, in which soil respiration was measured at individual
time points with non-automated chambers (Wertin et al., 2017). In that study,
the reduction with warming was observed 3 years after +2 ∘C
warming treatment was implemented. The dark respiration measurements were
made in the spring (at peak plant activity) and it was at the same point in
the season (see Fig. 2) that we saw the strongest seasonal driver for the
7-month cumulative data. In sum, although our NSE data do not allow us to
disentangle the driving mechanisms, changes in (i) biocrust composition,
(ii) nearby plant activity, and (iii) possibly surface moisture could all
have contributed to the reversal in the effect of the warming treatment in
the late period of the study. Regardless of the cause, these data suggest
large, sustained changes to dryland soil C cycling at our site in response to
climate change treatments.
We also observed reduced NSE values in the 2012–2013 sampling period in
plots that were previously watered plots compared to the control plots,
suggesting some legacy treatment effects. This was likely linked to loss of
mosses, cyanobacteria, or changes in vascular plant physiology. For example,
at a European site, biocrusted soil microsites were shown to be a dominant
source of midday soil respiration (Castillo-Monroy et al., 2011).
Furthermore, reductions in the autotrophic biomass seen with the climate
treatments could reduce respiration rates (Ferrenberg et al., 2017; Reed et
al., 2016). Plants accustomed to the extra water may also have responded
negatively to its absence, causing reduced physiological activity and hence
lower root respiration, an effect that has been documented in drought
simulation experiments (Talmon et al., 2011). Soil heterotrophs can also show
legacy effects of their species composition in response to changes in
precipitation regime (Kaisermann et al., 2017). Water retention may also have
been reduced due to the decline in biocrust cover, an effect for which there
is some evidence, particularly in semiarid ecosystems like our study site
(Belnap, 2006; Chamizo et al., 2012). Mosses have unique adaptations allowing
them to absorb high fractions of precipitation without loss to splash and
evaporation (Pan et al., 2016), a process that would be lessened in the
climate manipulation plots due to moss death. In addition to effects on soil
moisture, changes in biocrust community composition can have significant
effects on soil nutrient availability (Reed et al., 2012) and nutrient
availability can be tightly coupled with soil respiration rates (Reed et al.,
2011). Although the NSE data do not allow us to determine which gross C
fluxes caused the opposing treatment effects between the early (2006–2007)
and late (2012–2013) measurement periods, the observation of a reversal like
this is important because, if the larger CO2 loss had been sustained,
it would have indicated the potential for large feedbacks to increasing
atmospheric CO2 concentrations.
Interestingly, the CO2 loss reversal observed in 2013 did not
continue in 2014, likely due to the higher rainfall, particularly during
spring. In 2014, we saw high NSE in all plots in the 7-month cumulative
data, with no significant differences among treatments. Accompanying the
higher precipitation in 2014 – which occurred in a series of large rain
events in April and May – perennial plants were noticeably greener and there
was a flush of annual plants (Sasha C. Reed, unpublished data). During wet conditions, warmed plots had higher NSE
values, which could have been due to higher root respiration or higher
subsoil microbial activity, potentially linked to root turnover or
rhizodeposition (Jones et al., 2004). These results from the later period of
the experiment (2013–2014) underscore that taking a long-term perspective
(i.e., nearly a decade of warming) may be necessary for understanding climate
change effects, particularly those that maintain interactions with species
composition changes. Further, these data suggest more complexity in soil
CO2 efflux controls, such that some systems may not manifest a simple
transition from temperature-induced increases in soil CO2 loss to
temperature-induced decreases at later stages of warming. The interannual
variations in the magnitude of NSE fit with results from other drylands that
show high interannual variation in net ecosystem exchange (NEE) as measured
with eddy flux towers (Biederman et al., 2017). At least one other
longer-term manipulation in a dryland has also observed early stimulation of
plant growth with warming that then lessened over time, with longer-term
effects driven by changes in species composition (Wu et al., 2012). The
finding that decadal-scale studies can have mixed and context-dependent
effects not visible at the annual scale (Nielsen and Ball, 2015) is
exemplified in our study by the reversal in effects seen in 2013, followed by
the swamping out of those effects in a subsequent wet year.
Source of CO2 efflux
Observed NSE fluxes were almost always net positive (C loss to atmosphere),
indicating that soil profile C losses are greatly outpacing biocrust
photosynthetic uptake (Fig. 2). This necessitates a non-biocrust C source as
biocrusts cannot persist with consistently negative C balance (e.g., Coe et
al., 2012). The CO2 efflux data also support these non-biocrust
sources. For example, although we did lose biocrusts, even in control plots, C
losses continued even in plots where the larger biocrust constituents were
gone (e.g., watered plots in 2014). In addition to biocrust organisms, there are
three other potential sources of CO2 efflux: soil heterotrophs,
vascular plant roots, and pedogenic carbonates (Darrouzet-Nardi et al.,
2015). All three are possible contributors and further work is needed to
partition their contributions.
We would expect the biocrusts themselves to have the biggest impact on NSE
when soils are wet and biocrusts are active. During such time periods, we saw
treatment effects that were distinct from the 7-month totals (Fig. 2b),
which could be interpreted as evidence of a biocrust signal that did not
follow the general vascular plant trends of spring activity. Indeed, several
pieces of evidence point directly to a biocrust signal. First, in the later
time period (2013–2014), the reduction in minimum daily NSE during
precipitation events (Fig. 5) suggests that loss of biocrust CO2
uptake contributed to higher net C loss from these soils. In particular, the
combined treatment lost a large proportion of its capacity to
assimilate C, as well as much of the biocrust biomass. Second, the decline in
the size of the puffs of CO2 that were associated with the
1.2 mm watering treatments is likely driven by declines in biocrust
activity (Fig. 3), as these small watering events primarily affect the
surface of the soil. These biocrust activities could include both biocrust
respiration and decomposition of dead biocrust material. In our previous work
(Darrouzet-Nardi et al., 2015), we saw evidence of these puffs in control
plots without supplemental watering, though they were presumably not frequent
enough to kill the mosses under natural conditions, a situation that could be
altered if precipitation is altered in the future (Reed et al., 2012; Coe et
al., 2012).
Heterotrophic respiration could also be a substantial contributor to the
CO2 effluxes we observed. The soil CO2 efflux was observed
rapidly after each rain pulse (natural or experimental), which could indicate
soil heterotrophic respiration since plant photosynthesis may take longer to
become activated (López-Ballesteros et al., 2016). The soil organic C
pool in these soils includes ∼300 g C m-2 in the 0–2 cm
biocrust layer, which would be depleted rapidly if it were the sole C source.
However, the sub-biocrust 2–10 cm layer has ∼430 g m-2 and
soils are on average 50 cm deep at the site, suggesting that the total
sub-crust soil C is >1500 g C m-2 (data not shown). With a C pool
of that magnitude, depletion of soil organic matter C stocks could be
substantial contributors to the C losses we observed. However, if losses on
the order of 62 g C m-2 (the amount lost in control plots during
2006) were to continue, these stocks would be completely depleted (which
normally does not occur in soils) in ∼25 years, suggesting another
source is also extremely likely.
Root respiration is a contributor we consider highly likely. During
excavations of the chambers in 2017, root biomass was observed inside the
chambers, making a root signal plausible. Previously published measurements
from a nearby site that did not have a well-developed biocrust community
showed tightly coupled measurements of plant photosynthesis with soil
respiration directly beneath plant canopies (Wertin et al., 2015) while
correlations between soil C concentration and soil respiration were much
weaker (Wertin et al., 2018). Furthermore, the seasonal NSE trends are
broadly consistent with a plant photosynthetic signal, particularly the peak
in fluxes during the spring growing season, which coincides with plant uptake
as indicated by negative NEE seen using eddy flux towers (Darrouzet-Nardi et
al., 2015; Bowling et al., 2010). The interannual trends presented in this
study are also consistent with a plant signal: for example, the wettest year,
2014, was the year in which the highest CO2 efflux rates were
observed, a phenomenon that was likely driven by both increased activity in
perennials and the flush of annual plants observed in that year. Finally, not
only is a strong plant signal likely in these NSE measurements, but the
interpretation of the treatment differences, particularly the unexpected
finding of a reversal in the 7-month cumulative fluxes discussed above,
is clearer in light of a plant signal. We believe that by 2013, reductions in
plant productivity could have resulted in reduced root respiration in the
non-control plots.
Finally, pedogenic carbonates can contribute to CO2 efflux and we
cannot rule out their contribution in this study (Emmerich, 2003; Stevenson
and Verburg, 2006). Some studies suggest that CO2 efflux during dry
periods is likely to be from inorganic sources (Emmerich, 2003). Others make
the case that the timing of CO2 efflux from CaCO3 would be
more likely to overlap with the times when plants were active and calcite
could be dissolved in conjunction with a source of acidity such as acid
deposition, root exudation, or nitrification (Tamir et al., 2011). Either
way, long-term loss of CO2 from dissolved calcite from our site
cannot be ruled out and a field investigation of the isotopic composition of
released CO2 would be particularly valuable in assessing inorganic
contributions.
Conclusions
Both warming and watering with the associated moss death initially led to
higher CO2 losses in our experimental plots. After the cessation of
watering, the patterns in the C balances were reversed in an average moisture
year (2013), with the climate manipulation plots of all treatments showing
lowered soil CO2 loss relative to controls. These data are in line
with warming experiments from a range of climates, suggesting warming-induced
increases in soil CO2 are not a long-term phenomenon, at least within
these experimental frameworks. Moreover, in a subsequent wet year (2014),
CO2 fluxes were uniformly high among treatments. When focusing just
on periods of active biocrust photosynthesis, after 8 years, biocrust
photosynthetic performance was much weaker in both warmed and legacy watered
treatments relative to the control plots despite biocrust changes in control
plots as well. These results suggest that the community composition changes
that are highly likely in dryland plants (Collins et al., 2010; Wu et al.,
2011) and biocrusts (Ferrenberg et al., 2017; Johnson et al., 2012) as a
response to global change are likely to affect C balances even if effects are
not consistent year to year. Our results show how community shifts, such as
the loss of a major photosynthetic component like mosses, will contribute to
an altered C balance of these biocrusted soils. Finally, our results
underscore a strong role for biocrust, root, and possibly soil heterotrophic
and inorganic signals in NSE, suggesting that further study of the balance of
plant assimilation and root/rhizosphere respiration of C, as well as patterns
in biocrust C, in response to climate change will be an important determinant
of future C fluxes in drylands.