Fire Accelerates Assimilation and Transfer of Photosynthetic Carbon from Plants to Soil Microbes in...

Fire Accelerates Assimilationand Transfer of Photosynthetic

Carbon from Plants to Soil Microbesin a Northern Peatland

Susan E. Ward,1,2* Nick J. Ostle,2 Simon Oakley,2 Helen Quirk,1

Andrew Stott,2 Peter A. Henrys,2 W. Andrew Scott,2

and Richard D. Bardgett1

1Soil and Ecosystem Ecology Laboratory, Lancaster Environment Centre, Lancaster University, Bailrigg, Lancaster LA1 4YQ, UK;2Centre for Ecology and Hydrology, Lancaster Environment Centre, Library Avenue, Bailrigg, Lancaster LA1 4AP, UK

ABSTRACT

Northern peatlands are recognized as globally

important stores of terrestrial carbon (C), yet we

have limited understanding of how global changes,

including land use, affect C cycling processes in

these ecosystems. Making use of a long-term

(>50 year old) peatland land management exper-

iment in the UK, we investigated, using a 13CO2

pulse chase approach, how managed burning and

grazing influenced the short-term uptake and cy-

cling of C through the plant–soil system. We found

that burning affected the composition and growth

stage of the plant community, by substantially

reducing the abundance of mature ericoid dwarf-

shrubs. Burning also affected the structure of the

soil microbial community, measured using

phospholipid fatty acid analysis, by reducing fungal

biomass. There was no difference in net ecosystem

exchange of CO2, but burning was associated with

an increase in photosynthetic uptake of 13CO2 and

increased transfer of 13C to the soil microbial

community relative to unburned areas. In contrast,

grazing had no detectable effects on any measured

C cycling process. Our study provides new insight

into how changes in vegetation and soil microbial

communities arising from managed burning affect

peatland C cycling processes, by enhancing the

uptake of photosynthetic C and the transfer of C

belowground, whilst maintaining net ecosystem

exchange of CO2 at pre-burn levels.

Key words: peatland; burning; carbon cycle; sta-

ble isotope pulse labelling; 13C; respiration; photo-

synthesis; plant functional types; PLFA.

INTRODUCTION

Land use and land-use change for the production of

ecosystem goods and services affect around

one half of Earth’s land surface (Vitousek and

others 1997; Kareiva and others 2007), and is

widely recognized as the single most influential

determinant of global terrestrial carbon (C) stocks

(IPCC 2007; Ostle and others 2009). Northern

Received 1 March 2012; accepted 29 June 2012;

published online 9 August 2012

Electronic supplementary material: The online version of this article

(doi:10.1007/s10021-012-9581-8) contains supplementary material,

which is available to authorized users.

Author contributions: SEW, NJO, RDB conceived and designed the

study and wrote the paper. SEW, SO, HQ, AS performed the research and

contributed methods. SEW, PAH, WAS analysed data.

*Corresponding author; e-mail: [email protected]

Ecosystems (2012) 15: 1245–1257DOI: 10.1007/s10021-012-9581-8

� 2012 Springer Science+Business Media, LLC

1245

http://dx.doi.org/10.1007/s10021-012-9581-8

hemisphere peatland ecosystems are a globally

important store of terrestrial C, containing the

greatest organic C stocks of any terrestrial ecosys-

tem (Gorham 1991; Dise 2009); yet little is known

about how land management practices affect C

cycling processes in these ecosystems. Despite the

limited agricultural value of northern peatlands

(Heal and Smith 1978), these ecosystems have

been subjected to a long history of land-use change

in the UK, including forestry, peat extraction,

grazing and management for game which necessi-

tates regular managed burning (Ward and others

2007; Farage and others 2009). Well-managed

burns are relatively cool and short-lived, and are

designed to burn only the vegetation at intervals of

10–20 years, to create a mosaic habitat of old

and new plant shoots (Simmons 2003). This is in

contrast to wildfires, which cause extensive dam-

age to vegetation and soils, releasing considerable

amounts of C to the atmosphere (Turetsky and

others 2002). Managed burning at frequent inter-

vals, and wildfires over relatively longer-time

intervals, have both been shown to affect peatland

C dynamics. In Canada, changes in the peatland C

balance between sink and source have been

observed in the early years after wildfire, depend-

ing upon the vegetation community and whether

vegetation is sufficiently recovered for productivity

to exceed decomposition (Wieder and others 2009).

In addition, palaeo records in Canada show lower

peat accumulation with increasing fire frequency

(Kuhry 1994). In the UK, the peatland land man-

agement practice of managed burning and grazing

has also been shown to reduce ecosystem C stocks

(Garnett and others 2001; Ward and others 2007),

and alter C exchange as greenhouse gases (Ward

and others 2007; Clay and others 2010) and dis-

solved organic carbon (Worrall and others 2007;

Yallop and Clutterbuck 2009).

Burning and grazing dramatically affect the pro-

ductivity and composition of peatland vegetation

(Miles 1988; Rodwell 1991; Bardgett and others

1995), with potential impacts on ecosystem C

dynamics through changes in the quantity and

quality of C inputs to soil from plant litter and root

exudates, and C losses through decomposition and

respiration (Ward and others 2007; De Deyn and

others 2008; Hardie and others 2009). A change in

vegetation community composition is related to

contrasting plant functional traits, which dictate the

ability of each functional group to recover after dis-

turbances such as burning and grazing (Aerts and

others 1999; Ward and others 2009). In addition to

indirect effects of plant species, soil microbial com-

munity structure can also be directly affected by

physical disturbances such as fire (DeLuca and oth-

ers 2002). Many studies have established strong

links between peatland C cycling and abiotic condi-

tions of temperature and water table depth (Bubier

and others 1999; Updegraff and others 2001).

However, there is growing awareness that biotic

factors, such as changes in vegetation productivity

and composition, and changes in soil microbial

communities, also have an important impact on C

assimilation and losses in terrestrial ecosystems, with

feedbacks aboveground and belowground (Chapin

and others 1997; De Deyn and others 2008). Recent

studies using 13CO2 tracers have shown differences

in the assimilation and turnover of newly fixed C

among dominant plant functional groups in peat-

lands (Trinder and others 2008; Ward and others

2009) and in the high Arctic (Woodin and others

2009). These findings, combined with studies

showing variations in peatland carbon dioxide (CO2)

and methane (CH4) fluxes among peatland vegeta-

tion types (Heikkinen and others 2002; McNamara

and others 2008), indicate the importance of plant

community composition for short-term C cycling. It

is, however, unclear how fire- and grazing-induced

changes to vegetation and associated shifts in soil

microbial community structure affect peatland C

cycling processes.

In this study, we consider the effects of grazing

and managed vegetation burning (as opposed to

wildfire) on a European peatland ecosystem dom-

inated by low-lying vegetation. The overall aim of

our study was to quantify the effects of managed

fire- and grazing-induced changes on vegetation

composition and soil microbial communities on

short-term peatland C cycling, using a long-term

(>50 years) burning and grazing experiment in

northern England. More specifically, we used a

field-based 13C pulse-labelling approach (Ostle and

others 2000; Leake and others 2006) to test the

hypothesis that burning and grazing enhance the

photosynthetic uptake of 13CO2 by dominant plant

functional groups (ericoid dwarf-shrubs, grami-

noids and bryophytes), the transfer of this plant-

derived 13C to the soil microbial community, and

the return of recent photosynthetic 13C to the

atmosphere by respiration, thereby enhancing rates

of short-term C cycling processes through the

plant–soil system of northern peatland.

MATERIALS AND METHODS

Long-term Management Experiment

Experiments were performed on a long-term

study site, established in 1954, on an area of acidic

1246 S. E. Ward and others

ombrogenous blanket peat at Moor House National

Nature Reserve (NNR), in the North Pennines,

England, UK (54�65¢N, 2�45¢W; altitude 590 m).

Mean annual temperature is 6.1�C and mean

annual rainfall 2,012 mm, and peat is 1–2 m deep

(Ward and others 2007). Vegetation is low-lying,

and classified as a Calluna vulgaris-Eriophorum vag-

inatum blanket mire, Empetrum nigrum ssp nigrum

sub-community M19b (Rodwell 1991). This blan-

ket bog community, dominated by the ericoid

dwarf-shrub Calluna vulgaris, the sedge Eriophorum

vaginatum and hypnoid plus sphagnum mosses, is

typical of UK upland ombrogenous peat with harsh

winter conditions (Rodwell 1991). It is found

extensively in the north Pennines, Welsh moun-

tains and Scottish Highlands of the UK and, has

International equivalents in Europe (Eddy and

others 1969; Rodwell 1991; Simmons 2003). The

long-term burning and grazing land-use experi-

ment consists of four replicate blocks (60 9 90 m),

each with six 30 9 30 m plots. Within each block,

three plots are fenced off to exclude grazing.

Burning treatments of no burning, managed

burning at 10-year intervals and managed burning

at 20-year intervals are randomly allocated to

grazed and non-grazed areas (Figure 1), to give all

possible combinations of burning and grazing

treatments (Rawes and Hobbs 1979). This study

considers only 10-year interval burned plots, which

were last burned in winter 2007, and compares

them 18 months after burning with areas

unburned since 1954, for both grazed and ungrazed

treatments.

13CO2 Pulse Labelling13CO2 pulse labelling and subsequent sampling

were done in July 2008, approximately 18 months

after the most recent burning treatment. A plastic

base collar (30-cm diameter, 20-cm height) was

placed and surface sealed within each plot (depth of

5–10 cm) without cutting through roots at the soil

surface, and left to stabilize for a period of 10 days.

A clear chamber lid, 30-cm diameter and 35-cm

height (Ward and others 2007; Ward and others

2009), was then sealed on to the base collar using a

rubber band, and 99 atom % 13CO2 introduced

through a self-sealing septum in the chamber lid.

Each experimental plot was labelled with 40 ml of

99 atom % 13CO2 (Ward and others 2009) six times

between the hours of 11:00 and 15:00. During this

time, the mean level of photosynthetically active

radiation was 865 lmol s-1 m-2. Chamber lids

were left sealed for approximately 30 min after

each addition of 13CO2, and then removed for

10 min between pulses to allow the vegetation and

Figure 1. Experimental

site aerial photograph and

schematic diagram.

Fire Accelerates Peatland Carbon Cycling 1247

chamber headspace to re-equilibrate with the

atmosphere. Plant and soil samples and respired

ecosystem CO2 were collected immediately fol-

lowing 13CO2 pulse labelling and after 1, 3, 8, 15

and 22 days to trace the assimilation and retention

of recently fixed photoassimilate C.

Vegetation and Soil Sampling

Live photosynthetically active plant shoot material

was sampled for all species present within the 13C-

labelled plots. This comprised approximately 2-cm

lengths for dwarf-shrubs, 2–3 whole leaves for

graminoids, and the top 2–3 cm of bryophytes and

lichens. Small-sized soil samples (�2.5 9 2.5 cm,

0–10-cm depth) were collected from inside the

pulse-labelled base collar on each sampling occasion,

and care was taken to minimize disturbance and

avoid the creation of gaps in the soil surface. Vege-

tation and soil samples were frozen and subse-

quently freeze dried and ground. Dried vegetation

and soil samples were analyzed for 13C at the NERC

Life Sciences Mass Spectrometer Facility, CEH Lan-

caster, using a Carbo Erba elemental analyzer linked

to a Dennis Leigh Technologies Isotope Ratio Mass

Spectrometer (IRMS). All vegetation analyses were

done by individual species, and then combined to

present results at a plant functional group level. At

the end of the experiment, 22 days after application

of the 13CO2 pulse, all aboveground vegetation

within the pulse-labelled area was destructively

harvested, and material oven dried to determine

biomass dry weights.

Phospholipid fatty acids (PLFAs) were extracted

from 0.3 g of freeze-dried soils collected on the first

four days of sampling, using the Bligh-Dyer (White

and others 1979) extraction method (2004). PLFAs

were quantified by gas chromatography using Ag-

ilent Technologies 5973 Mass Selective Detector

coupled to Agilent Technologies 6890 GC, with

concentrations calculated for all identifiable PLFAs

via an internal standard method (C19 FAME Sigma

Aldrich). The fatty acids i15:0, a15:0, 15:0, i16:0,

16:1x7, 16:1, 16:1x5, i17:0, a17:0, cy17:0 18:1x7,

18:1x5 and cy19:0 were taken to represent bacte-

rial PLFAs, and 18:2x6 as fungal PLFA marker

(Harrison and Bardgett 2010). The ratio of 18:2x6

fungal marker to bacterial PLFAs was taken to

represent the ratio of fungal-to-bacterial biomass in

soil (Bardgett and others 1996). PLFAs were ana-

lyzed for 13C by gas chromatography-combustion-

isotope ratio mass spectrometry (GC-C-IRMS) at

the NERC Life Sciences Mass Spectrometry Facility,

CEH Lancaster. d13C values were measured on

individual PLFAs using an Isoprime isotope ratio

mass spectrometer interfaced via a combustion

furnace to Agilent 6890 GC.

Trace Gas Flux Measurements

Ecosystem respiration was measured between the

hours of 12:00 and 14:00 on the same day as vege-

tation and soil sampling, using the same chamber lids

as for the isotope pulse application with the addition

of an opaque black-out bag to prevent photosyn-

thesis (Ward and others 2007). Gases were sampled

when the chamber was first sealed and at two sub-

sequent time points over an average 21 min of clo-

sure. This sampling schedule was based on previous

experience at this site which showed that fluxes

reliably increased in a linear way over this headspace

closure time (Ward and others 2009). Samples were

taken through the septum using a 20-ml syringe

fitted with a 0.5-mm gauge needle, and transferred

into evacuated exetainers (Labco Ltd, UK) before

laboratory analysis. Respired CO2 samples were

analyzed for 13C at the NERC Life Sciences Mass

Spectrometer Facility, CEH Lancaster, using a Trace-

gas Preconcentrator coupled to an Isoprime Isotope

Ratio Mass Spectrometer (Isoprime Ltd). Additional

measurements of net CO2, photosynthesis, and CH4

fluxes were made on the same sampling days using a

transparent and dark chamber technique (Wadding-

ton and Roulet 2000; Nykanen and others 2003;

Ward and others 2007), and rates of photosynthesis

were calculated by the difference between transpar-

ent (net) flux and dark (respiration) flux. Air tem-

perature, soil temperature and photosynthetically

active radiation were recorded during all sampling

events. Samples were analyzed for CO2 and CH4 by

gas chromatography (GC) within 6 weeks of collec-

tion, on a Perkin Elmer Autosystem XL GC with

Flame Ionization Detector containing a methanizer,

calibrated against certified gas standards (Air Prod-

ucts, UK). All fluxes were adjusted for field sampling

temperature, headspace volume and chamber area

(Holland and others 1999), and calculated using a line

of best fit between the three time points sampled.

Isotopic Mass Balance and StatisticalAnalysis

Results for 13C enrichment in vegetation and

respired CO2 are reported as 13C atom % excess, in

line with convention for samples highly enriched

with 13C following tracer application (Boutton

1991). Atom % excess is calculated as follows:

atom % excess = atom %enriched sample - atom

%background sample, where atom % = [Rsample/

(Rsample + 1)] 9 100, and where Rsample is the


13C/12C ratio determined by IRMS. For vegetation

samples, background values were calculated from

vegetation tissues collected before pulse labelling

with 13CO2. For respired CO2 samples, background

values were from headspace gas samples taken at

t0. Pulse-derived 13C per m2 of peatland was cal-

culated from the 13C enrichment of each sample

multiplied by the C content of each sample; in

vegetation as mg C per g shoot biomass, and for gas

flux as mg CO2–C h-1 in respired CO2 (Trinder and

others 2008). Results for 13C in bulk soil and PLFA,

which were not so highly enriched in 13C, are

reported as d13C &.

Statistical analyses were done using SAS Enter-

prise Guide 4. Data were checked for normality

using residual plots method, and log transformed

when necessary before analysis. Data from the final

vegetation harvest were analyzed by ANOVA using

generalized linear models to ascertain the effects of

burning and grazing, and any interactions between

them. All other data were analyzed by repeated

measures of ANOVA. For the shoot tissue 13C

enrichment over the 22-day pulse-labelling period,

data for each species/treatment combination were

well approximated by an exponential decay curve

tending to a lower limit rather than to zero. A

three-parameter curve: y =a + (y - a) * (1 - bs),

was fitted to the 13C enrichment measurements of

each species/treatment/plot combination. The var-

iable y represents 13C enrichment at time t after the

pulse labelling. The parameters are interpretable as

a, the initial uptake of 13C on the day of pulse label;

b, a measurement of the rate at which 13C

enrichment was lost over the 22-day pulse chase

period; and c the lower limit to the fitted curve, or

approximately the amount of 13C remaining in

plant shoot tissue at day 22. The set of estimates for

each of these parameters was then analyzed by

ANOVA to examine grazing, burning and species

effects. Finally, after first checking for linearity of

the data, and low common absences (<2% of all

data), principal component analysis (PCA) for all

detected individual PLFAs was used to assess

microbial community differences and 13C enrich-

ment between burning and grazing treatments.

RESULTS

Vegetation and Soil MicrobialCommunity Composition

Burning reduced total aboveground live vegetation

biomass by over 70% (F1,16 = 20.1, P = 0.001), as a

consequence of a substantial reduction in dwarf-

shrubs. This altered the relative contribution of the

three plant functional groups by increasing the

proportion of total aboveground biomass repre-

sented by graminoids and bryophytes/lichens

(Table 1). In contrast, grazing did not affect vege-

tation biomass (F1,16 = 0.9, P = 0.37), and there

was no interaction detected between burning and

grazing (F1,16 = 1.0, P = 0.34).

In the soil microbial community, burning reduced

the total abundance of fungal PLFA to less than a

quarter of the value in unburned soils (F1,63 = 122.7,

P < 0.0001), but had no detectable effect on bac-

terial PLFA (F1,63 = 0.0, P = 0.94) or the total

abundance of PLFA (F1,63 = 0.4, P = 0.52) (Table 2).

As a consequence, the ratio of fungal:bacterial PLFA

was reduced from 0.41 (±0.02) to 0.07 (±0.01) by

burning (F1,63 = 93.0, P < 0001). Grazing had no

detectable effect on fungal PLFA (F1,63 = 1.3,

P = 0.27), bacterial PLFA (F1,63 = 2.0, P = 0.18),

total PLFA (F1,63 = 2.9, P = 0.12) or fungal:bacterial

PLFA ratio (F1,63 = 0.8, P = 0.41), although there

was a trend towards lesser total abundance of PLFA

due to grazing. Differences in soil microbial PLFA

abundance due to burning and grazing are further

evidenced by PCA analysis of the relative abundance

of the 23 individual PLFAs (Figure 2). Principal

component axis 1 primarily reflects differences

because of burning and explained 62.3% of the

Table 1. Live Aboveground Vegetation Biomass in Unburned and Burned Areas

Plant functional group Unburned Burned

Dry weight

(g m-2)

% of total

vegetation

Dry weight

(g m-2)

% of total

vegetation

Dwarf-shrubs (photosynthetic tissues) 192.2 (±33.1) 22.29 15.8 (±11.2) 6.40

Dwarf-shrubs (stems) 468.4 (±45.7) 54.31 13.1 (±9.5) 5.29

Graminoids 49.8 (±14.1) 5.78 62.6 (±5.2) 25.26

Bryophytes and lichens 151.9 (±65.1) 17.62 156.2 (±56.0) 63.05

Total live vegetation 862.4 (±76.5) 100.00 247.7 (±73.2) 100.00

Values are in g dry weight m-2 (±SE) and percentage contribution to total vegetation, for each of the three plant functional groups. (n = 8 unburned and 8 burned).


variation in the data (P = 0.001), whereas principal

component axis 2 explained 21.5% of the variation

in data and, although weak (P = 0.548), is likely to

be related to the grazing treatment. No significant

interactions between burning and grazing were

detected for fungal (F1,63 = 0.9, P = 0.36), bacterial

(F1,63 = 0.9, P = 0.37), or total PLFA(F1,63 = 2.7,

P = 0.13).

Trace Gas Fluxes—CO2 and CH4

There were no significant differences in gross and

net CO2 fluxes due to either the burning or grazing

treatments (Table 3, Table 4). Fluxes of CO2 did

vary between sampling dates, because of differ-

ences in temperature and solar radiation, correlat-

ing most strongly to daytime air temperatures (R2

0.62, 0.34 and 0.37 for respiration, photosynthesis

and net flux, respectively) which ranged from 11.2

to 18.3�C, with a mean of 14.9�C. There were also

no significant differences in CH4 flux due to

burning (F1,72 = 2.6, P = 0.14), grazing (F1,72 =

1.2, P = 0.29) or sampling date (F5,72 = 0.6

P = 0.72). There was, however, a trend for higher

CH4 flux in areas that had been burned relative to

unburned areas, at 1.9 (±0.2) mg m-2 h-1 com-

pared with 1.4 (±0.2) mg m-2 h-1 (Tables 3, 4).

Vegetation Assimilation and Turnoverof 13C Tracer

The level of 13C assimilation by shoot tissues on the

initial day of pulse labelling represents the amount

Table 2. Soil Total PLFA Concentration (n moles g dry weight soil-1), in Unburned Versus Burned andUngrazed Versus Grazed Areas

Soil PLFA concentration (n moles g dry weight soil-1)

Unburned Burned Ungrazed Grazed

Total PLFA 1113.4 (±41.6) 1032.7 (±74.1) 1179.9 (±47.2) 977.9 (±63.9)

Fungal PLFA 235.5 (±12.0)a 40.7 (±6.2)b 152.7 (±19.0) 127.8 (±21.1)

Bacterial PLFA 584.7 (±26.9) 583.6 (±41.3) 636.8 (±30.2) 536.4 (±35.7)

Fungal: bacterial PLFA ratio 0.41 (±0.02)a 0.07 (±0.01)b 9.5 (±2.1) 15.1 (±2.6)

Values are means of all sampling dates (±SE). Different letters indicate significant difference between burning treatments in the row.

Figure 2. PCA analysis of PLFA abundance for burning

and grazing treatments. Data are from 23 PLFAs (n moles

g dry weight soil-1), for four dates after 13CO2 pulse

labelling (n = 64). Axis 1 = 62.3%, Axis 2 = 21.5.

Table 3. Trace Gas Fluxes (CO2, CH4 in mg m-2 h-1) in Unburned and Burned Areas

Flux (mg m-2 h-1)

Unburned Burned

Ecosystem respiration 498.2 (±33.4) 459.6 (±29.6)

Gross primary productivity -441.0 (±50.8) -429.2 (±32.8)

Net ecosystem exchange (CO2) 35.9 (±13.9) 16.60 (±24.0)

Methane flux 1.37 (±0.23) 1.94 (±0.18)

Values are means of all sampling days (±SE).


of newly fixed 13CO2, less any 13C used and

translocated during the 4-h pulse-labelling period.

We found an interaction between burning and

plant functional groups (F2,59 = 10.8, P = 0.0001),

highlighting a difference in the amount of newly

fixed 13C between the three plant functional groups

after burning (Table 5, Figure 3A, B, C). Dwarf-

shrubs from burned areas showed over twice as

much 13C enrichment on the day of pulse labelling

relative to unburned areas (F1,20 = 7.6, P = 0.02)

(Figure 3A), graminoids showed little difference

(F1,19 = 2.9, P = 0.11) (Figure 3B), whereas a 20-

fold enrichment in 13C was detected in photosyn-

thetic tissues of bryophytes from burned areas

(F1,20 = 15.4, P = 0.002) (Figure 3C). In contrast,

grazing had no detectable effect on shoot 13C

enrichment following the 13CO2 pulse (F1,59 = 1.4,

P = 0.25). The mean loss of 13C enrichment in

plant shoot tissues as the 13C pulse was diluted with

time was 54% by day three and 74% by day 15.

The rate of loss of newly fixed 13C over the 22 days

differed between plant functional groups

(F2,59 = 10.2, P = 0.0002), but was unaffected by

burning (F1,59 = 0.2, P = 0.66) or grazing

(F1,59 = 0.8, P = 0.39) (Table 5). By the end of the

pulse chase period on day 22, only bryophytes still

showed significantly greater 13C enrichment in

burned relative to unburned treatments

(F1,59 = 21.6, P < 0.0001). Total pulse-derived 13C

in shoot tissues, calculated per dry weight of

aboveground plant biomass present over a m2 area

of peatland, was 2.5 times greater in burned rela-

tive to unburned areas over the 22 pulse chase

period, with the greatest enrichment (3.6 times) on

the day of pulse labelling. (F5,96 = 2.8, P = 0.026)

(Figure 4A).

Differences in shoot tissue 13C enrichment at a

species level within the functional groups were also

detected on the day of pulse labelling for ericoid

dwarf-shrubs (F3,20 = 3.4, P = 0.04) and bryo-

phytes (F8,20 = 3.3, P = 0.04), although these

findings need to be treated with caution because of

the low number of repetitions of each species

present within the sampling plots. For dwarf-

shrubs, the dominant species Calluna vulgaris was

30% more enriched with 13C than other dwarf-

shrub species (Empetrum and Vaccinium sp.). For

bryophytes, the acrocarpous mosses (Polytrichum

Aulacomnium and Dicranum sp.) showed 20 times

greater enrichment than the pleurocarpous mosses

(Hypnum, Pleurozium and Plagiothecium sp.). No

difference in shoot tissue 13C enrichment was

Table 4. Effects of Sampling Date, SamplingBlock, Burning and Grazing on Trace Gas Fluxes ofCO2 and CH4 in mg m-2 h-1

Source of variation df Flux

(mg m-2 h-1)

F P

Respiration

Sampling date 5 13.6 <0.0001

Sampling block 2 13.1 0.0006

Burning 1 1.2 0.30

Grazing 1 0.7 0.43

Burning 9 sampling date 5 1.1 0.36

Estimated photosynthesis

Sampling date 3 5.6 0.004


Burning 1 0.0 0.87

Grazing 1 0.4 0.55


Net CO2 flux



Burning 1 0.5 0.49

Grazing 1 0.2 0.70


CH4



Burning 1 2.6 0.14

Grazing 1 1.2 0.29


P values in bold indicate significant difference (P < 0.05).

Table 5. Effects of Sampling Block, Plant Func-tional Group, Burning and Grazing on 13C Enrich-ment of Plant Shoot Tissue

Source of variation df F P

(a) Initial uptake of 13C (day of pulse label)


Plant functional group 2 118.1 <0.0001

Burning 1 29.4 <0.0001

Grazing 1 1.4 0.25

Plant functional group 9 burning 2 10.8 0.0001

(b) Rate of loss of 13C (over 22 day period)


Plant functional group 2 10.2 0.0002

Burning 1 0.2 0.66

Grazing 1 0.8 0.39

Plant functional group 9 burning 2 1.0 0.39

(c) 13C remaining 22 days after pulse label


Plant functional group 2 98.1 <0.0001

Burning 1 21.6 <0.0001

Grazing 1 1.0 0.32

Plant functional group 9 burning 2 12.3 <0.0001

P values in bold indicate significant difference (P < 0.05).


observed between the two graminoid species

(F1,19 = 0.26, P = 0.62), Eriophorum vaginatum and

Eriophorum angustifolium.

Soil and Soil Microbial Incorporationof 13C Tracer

The mean natural abundance d13C of the soil was

-27.35& (±0.04 SE), typical of C3 soil, and there

was no difference in natural abundance d13C due to

either burning (F1,16 = 2.3, P = 0.16) or grazing

(F1,16 = 0.04, P = 0.85). After application of the 13C

pulse, soils from burned areas were found to be

slightly enriched in 13C at -27.06& (±0.08 SE)

compared with -27.61& (±0.12 SE) for unburned

soils (F1,80 = 8.6, P = 0.01). Neither grazing (F1,80 =

0.07, P = 0.80) nor sampling date (F4,80 = 0.4,

P = 0.85) affected soil 13C enrichment.

To assess the uptake of the 13C label by soil

microbial communities, d13C values were obtained

for 17 detectable PLFAs, before pulse labelling, on

the day of pulse labelling and 1, 3 and 8 days after

application of the 13CO2 pulse. Mean d13C values

showed greater 13C enrichment in burned plots for

16 out of 17 PLFAs on the day of pulse labelling and

14 out of 17 PLFAs for 1, 3 and 8 days after pulse

labelling, relative to unburned plots (supplemen-

tary information). PLFA d13C values show cluster-

ing of both burning and grazing treatments in the

PCA using all 17 PLFA d13C values for all sampling

days (Figure 5). Axis 1 of the PCA analysis explains

53.7% of the variation in the data, indicating a

strong difference in the uptake of newly photo-

synthesized C between microbial communities

present in burned and unburned areas. Principal

component axis 2 explained 26.8% of the variation

in data and is likely to be related to the grazing

treatment. PLFA d13C values are strongly correlated

with plant shoot tissue (correlation: 0.35; P =

0.008) and soil (correlation: 0.34; P = 0.009) 13C

enrichment.

Ecosystem Respiration of 13C Tracer

After application of the 13CO2 pulse, 13C enrichment

was detected in respired CO2 on all sampling days.

Levels of enrichment declined exponentially over

Figure 3. 13C enrichment in plant shoot tissues for

burned and unburned treatments. Data are for up to

22 days after 13CO2 pulse labelling, for: A dwarf-shrubs;

B graminoids; C bryophytes and lichens (13C atom %

excess). Figures are means ± SE.

0.0

1.0

2.0

3.0

4.0

Pul

se d

eriv

ed13

C in

sho

ot ti

ssue

s (m

g g

dry

wei

ght-1

m-2

)

A

Burned

Unburned

0.0

2.0

4.0

6.0

8.0

10.0

0 5 10 15 20 25Pul

se d

eriv

ed13

C in

res

pire

d C

O2

-C

(mg

m-2

hr-1

)Days after 13CO2 pulse application

B

Burned

Unburned

Figure 4. Pulse-derived 13C for burned and unburned

treatments. Data are for up to 22 days after 13CO2 pulse

labelling, for: A in plant shoot tissues (mg g dry weight-1

m-2); B in respired CO2 (mg m-2 h-1). Figures are

means ± SE.


the 22-day pulse chase period as the initial pulse was

diluted (F5,71 = 295, P < 0.0001), with 83% having

been lost by day three, and 98% by day 15. There was

no significant difference in 13C enrichment in

respired CO2 due to either burning (F1,71 = 0.05,

P = 0.83) or grazing (F1,71 = 0.02, P = 0.90)

(Figure 6). Total pulse-derived 13C returned to the

atmosphere in respired CO2 (mg m-2 h-1) also did

not differ between burned and unburned areas

(F1,70 = 0.11, P = 0.75) (Figure 4B).

DISCUSSION

The aim of our study was to determine the effects of

land management-induced changes in vegetation

composition and soil microbial communities on

short-term peatland C cycling on a peatland burn-

ing and grazing experiment. Our findings demon-

strate that, in line with our hypothesis, burning

significantly enhanced rates of photosynthetic

assimilation of CO2 and the transfer of the newly

fixed C into soil microbial communities through

the plant–soil system. This was associated with

changes in plant community composition and

growth stage, and changes in soil microbial com-

munity structure. More specifically, we found that

burning reduced the biomass and relative contri-

bution of mature dwarf-shrubs aboveground, the

biomass of soil fungi, and the fungal:bacterial ratio

of the soil microbial community. Increased photo-

synthetic assimilation of 13C in burned areas can be

attributed to greater assimilation of 13C by shoot

tissues of early growth stages of dwarf-shrubs and

bryophytes as they recover from burning, which

was accompanied by reduced soil fungal biomass

and a greater transfer of photosynthetic C into the

microbial community. Grazing, which is known to

affect soil nutrient and C cycling through a variety

of mechanisms (Bardgett and Wardle 2003), had no

detectable effects on short-term C cycling and plant

community composition, and only minor effects on

the soil microbial community, despite this treat-

ment being imposed for more than 50 years. This is

most likely due to the low grazing pressures on this

moorland system (summer grazing only, at

<1 sheep ha-1) and consequent lack of effects of

grazers on vegetation composition and nutrient

cycling via selective grazing, the return of feces and

physical disturbance.

We detected no difference in the rates of

exchange of C with the atmosphere as gross and net

CO2 fluxes, or as CH4, between burned and un-

burned areas. This was despite observed changes in C

assimilation and plant–soil C cycling, a reduction in

plant aboveground biomass, and a change in soil

microbial community structure, measured as the

fungal:bacterial ratio, due to burning. Vegetation

has been estimated to account for around 30%

(Dorrepaal and others 2009) and up to 54% (Hardie

and others 2009) of total ecosystem respiration in

peatlands. Based on these figures, and a mean un-

burned respiration rate of 498 mg CO2 m-2 h-1

measured in this study (Table 3), the expected mean

rates of respiration for burned areas with 70% less

vegetation (Table 1) would be 393 and 310 mg

CO2 m-2 h-1, for the lower and higher respiration

estimates, respectively. In contrast, our measured

mean flux for ecosystem respiration in burned areas

is 460 mg m-2 h-1. Similarly, owing to the sub-

stantially lower vegetation biomass in burned areas,

we would have expected to see a similar reduction in

gross primary productivity (Table 3), calculated

Figure 5. PCA analysis of 13C enrichment in PLFAs, for

burning and grazing treatments. Data are from 17 PLFAs

(d13C &) for four dates after 13CO2 pulse labelling

(n = 64). Axis 1 = 53.7%, Axis 2 = 26.8%.

Figure 6. 13C enrichment in respired CO2 for burned

and unburned treatments. Data are for up to 22 days

after 13CO2 pulse labelling (13C atom % excess). Figures

are means ± SE.


from the difference between ecosystem respiration

and net ecosystem exchange. These extrapolated

figures need to be treated with some caution, be-

cause of uncertainties in the complex relationships

between vegetation biomass, soil microbial com-

munities and CO2 fluxes, and because our sampling

was done over only a short period of time. Never-

theless, they suggest greater rates of gross respiration

and photosynthesis in burned compared with un-

burned areas of this peatland ecosystem.

Previous study at this field site, carried out

9 years after a burn when shoot biomass in burned

areas had recovered to around 50% of unburned

values (Ward and others 2007), found that rates of

respiration and photosynthesis were greater in

burned relative to unburned areas. However, at

this time, and unlike in this study, burning in-

creased the net sink of CO2 because of a greater

accelerating effect of burning on photosynthesis

relative to respiration (Ward and others 2007). A

similar link between vegetation composition and

biomass with C sink/source function was recently

observed in a peatland wildfire chronosequence in

Canada (Wieder and others 2009), where changes

in C dynamics were attributed to recovery of veg-

etation following the disturbance, with peak C sink

strength estimated at 75 years after fire. Variations

in ecosystem process rates with time since burning

have also been shown in other studies, for soil

respiration (Amiro and others 2003; Hubbard and

others 2004; Michelsen and others 2004) and soil

microbial properties (DeLuca and Zouhar 2000;

Hart and others 2005). This growing body of evi-

dence highlights the need to consider changes in

plant functional group composition and ecosystem

C fluxes over a range of time scales after a burn

event, to calculate the long-term effect of fire on

the peatland C cycling.

At a plant functional group level, evidence for

enhanced rates of C assimilation after burning

comes from observations of 13C enrichment in

plant shoot tissues. Our findings show that the

amount of 13CO2 taken up by dwarf-shrubs and

bryophytes was greater after burning, indicating

that plant growth stage (that is, recovery time since

burning) has a substantial effect on the assimilation

of new C, and that the strength of this effect is

dependent upon plant functional group identity.

We also observed differences between the three

vegetation groups in the rate at which the 13CO2

pulse label was assimilated, with dwarf-shrub and

graminoid shoot tissues taking up nearly ten times

more 13C than bryophytes and lichens (Figure 3A,

B, C), supporting the findings of earlier peatland

pulse-labelling studies (Trinder and others 2008;

Ward and others 2009). We calculated that the

total pulse-derived C per g of plant shoot tissue

present in a square meter of peatland on the day of

pulse labelling was 3.6 times greater in the burned

compared with unburned areas, providing further

evidence of a relatively greater rate of photosyn-

thetic uptake of C by vegetation after burning. For

dwarf-shrubs, the photosynthetic tissues of the

younger, smaller plants recovering from burning

were able to assimilate twice as much 13C relative

to the photosynthetic tissues of the older shrubs in

unburned plots, showing a more active uptake of

new C in the early stages of recovery from burning.

The greatest proportional difference in 13C shoot

tissue enrichment between burned and unburned

areas was observed for bryophytes and lichens, al-

though as stated earlier, the contribution of bryo-

phytes to overall pulse-derived 13C is relatively

small due to the magnitude of 13C uptake being

substantially lower than dwarf-shrubs and grami-

noids. The 20-fold increase in 13C assimilation

observed in bryophyte and lichen tissues from

burned relative to unburned areas may have been

due to removal of the dwarf-shrub canopy and a

consequent increase in available solar radiation

(Grace and Marks 1978). This increase observed

could also be attributed to a change in species com-

position as greater levels of 13C assimilation were

observed in acrocarpous moss species, which are

typically abundant in early successional communi-

ties following disturbances such as burning, con-

trasting with the later successional pleurocarpous

species which had lower 13C tissue enrichment.

Differences in C cycling process rates between

bryophyte species were also observed by Lang and

others (2009), who found significantly lower

decomposition rates in Sphagnum compared with

non-Sphagnum mosses and liverworts. These find-

ings collectively highlight the potential for species-

level differences in the impact of bryophytes to

influence rates of C cycling in peatland. In addition,

the lower plant functional group retained a greater

amount of newly photosynthesized 13C in shoot

tissues at the end of the 22-day pulse chase period

compared with dwarf-shrubs and graminoids. This

implies that, despite low rates of photosynthetic C

uptake, mosses and lichen play an important role in

long-term ecosystem C storage, a finding echoed by

Woodin and others (2009) in a tracer study in the

high Arctic. In contrast, graminoids showed no

difference in shoot tissue 13C enrichment between

burned and unburned treatments. This might be

explained by the fact that graminoids are senescent

during the winter when prescribed moorland

burning is permitted, making them less vulnerable


to tissue damage. However, the presence of Erio-

phorum has been linked to enhanced CH4 emissions

because of the presence of aerenchaemous tissues

(Greenup and others 2000), and this may explain

the greater CH4 fluxes observed in burned relative

to unburned areas.

The increase in rates of transfer of the newly

fixed C cycling across the plant–soil interface in

burned relative to unburned areas evidenced from

our findings can be related to differences in soil

microbial community composition, measured as

PLFAs, and the rate of enrichment in 13C in PLFAs.

The reduction in fungal PLFA in burned areas to

less than a quarter of the level in unburned areas

could be due to the physical loss of material in litter

and soil F and H horizons following the burn (Ward

and others 2007), to the greater vulnerability of

decomposer fungi to direct effects of fire and heat

compared with bacteria (Hart and others 2005) or

to loss of mycorrhizal fungi associated with ericoid

dwarf-shrubs. It could also potentially be a conse-

quence of enhanced input of more labile C from the

faster growing plants after burning, with a reduced

fungal:bacterial ratio being typically associated

with elevated rates of decomposition and nutrient

cycling (van der Heijden and others 2008; Bardgett

and Wardle 2010). Indeed, it has been proposed

that fungal-based energy channels are associated

with ‘slow’ nutrient cycling and bacterial-based

energy channels with ‘fast’ nutrient cycling (War-

dle and others 2004; Bardgett and Wardle 2010),

providing a possible explanation for the accelera-

tion of C cycling processes in burned areas with

reduced fungal-to-bacterial PLFA ratios. The

greater levels of 13C enrichment in bulk soil and

PLFAs indicate a more rapid transfer of newly

photosynthesized C belowground in burned rela-

tive to unburned areas. Such rapid transfer of new

photosynthates belowground into soil microbial

communities has been shown in other 13C pulse-

labelling studies (Ostle and others 2003; Jin and

Evans 2010; De Deyn and others 2011) and adds to

the growing body of evidence that recent photo-

synthates act as an important driver of ecosystem C

dynamics (Hogberg and Read 2006).

Overall, our study presents evidence that man-

aged fire accelerates the assimilation and transfer of

photosynthetic C within the plant–soil system of a

peatland ecosystem, but has no effect on net eco-

system CO2 exchange. In contrast, long-term

grazing, albeit at low grazing pressures, had no

detectable effects on any of the C cycling processes

measured. We propose that the changes in C

assimilation and transfer of C to soil microbes

observed during the summer in the recently

burned treatment plots can be attributed to changes

in the functional composition and the age of the

plant community post-burning, and to changes in

soil microbial community. Gross CO2 flux measures

of ecosystem respiration and gross primary pro-

ductivity also indicate enhanced gross rates of C

cycling in burned relative to unburned areas, when

the substantial difference in photosynthetic bio-

mass is accounted for. It is not possible to deter-

mine, however, what proportion of changes are

due to increased/reduced photosynthesis or chan-

ges in soil microbial activity. This study, which took

place over a 3-week period during the summer

growing season, provides a snap-shot of peatland C

cycling processes 18 months after a managed burn

in a peatland dominated by low-lying vegetation.

As such, it does not attempt to calculate compara-

tive burn/no burn C budgets for this peatland

ecosystem, nor does it account for physical C losses

from the actual burn. What our findings do provide

is a new insight into how changes in vegetation and

soil communities arising from managed burning

affect peatland C cycling processes, by enhancing

the uptake of photosynthetic C and the transfer of

C belowground, whilst maintaining net ecosystem

exchange of CO2 at pre-burn levels. Although these

findings are most directly applicable to managed

peatlands in the UK, we suggest that this new

mechanistic understanding of how changes in plant

functional diversity affect carbon cycling processes

are potentially transferrable to other northern

peatlands.

ACKNOWLEDGEMENTS

This research was supported by the Natural Envi-

ronment Research Council (NERC) EHFI grant

(NE/E011594/1) awarded to R. D. Bardgett and N.

J. Ostle. The authors thank Hannah Tobermann

and Emily Bottoms for their help in the field, Helen

Grant for her stable isotope analysis, and two

anonymous referees for their helpful comments on

an earlier versions of this manuscript. The authors

also thank Natural England and the Environmental

Change Network, CEH, Lancaster for access to and

information on the field site.

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