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Soil Biology & Biochemistry 38 (2006) 20112024

Decomposition in peatlands: Reconciling seemingly contrasting resultson the impacts of lowered water levels

Raija Laiho

Peatland Ecology Group, Department of Forest Ecology, University of Helsinki, P.O. Box 27, FIN-00014 Helsinki, Finland

Received 9 September 2005; received in revised form 3 February 2006; accepted 22 February 2006

Available online 4 April 2006

Abstract

Northern peatlands represent about 30% of the global soil C pools. The C pool in peat is a result of a relatively small imbalance

between production and decay. High water levels and the consequent anoxia are considered the major causes for the imbalance. As such,

the C sink of a peatland is labile, and sensitive to disturbances in environmental conditions.

Changes in peatland ecosystem functions may be mediated through land-use change, and/or climatic warming. In both cases, lowering

of the water level may be the key factor. Logically, lowered water levels with the consequent increase in oxygen availability in the surface

soil may be assumed to result in accelerated rates of organic matter decomposition. Yet, earlier research has given highly contrasting

results concerning the effects of lowered water levels on the rates of decomposition and the C sink/source behaviour of peatlands. The

mechanisms controlling this variation remain unresolved.

This paper summarizes the changes observed in the biotic and abiotic controls of decomposition following natural or artificial lowering

of peatland water levels and show that they are complex and their interactions have not been previously explored. Long-term changes in

the C cycle may differ from short-term changes. Short-term changes represent a disturbance in the ecosystem adapted to the pre-water-

level-lowering conditions, while long-term changes result from several adaptive mechanisms of the ecosystem to the new hydrological

regime. While in a short term, the disturbed system will always lose C, the long-term changes inherently vary among peatland types,

climates, and extents of change in the water level. The paper closes by identifying the gaps in our knowledge that need to be addressed

when proceeding towards a causal and unifying explanation for the C sink/source behaviour of peatlands following persistent lowering of

the water level.

r 2006 Elsevier Ltd. All rights reserved.

Keywords: C cycling; Decomposition dynamics; Decomposer communities; Hydrological change; Litter quality; Long-term change; Secondary succession;

Short-term change; Wetland ecology

1. Introduction

Peatlands represent a wide variety of wetlands that are

characterized by an organic soil, but differ in hydrology,chemistry, and, consequently, vegetation composition.

Northern peatlands have been a significant sink of carbon

(C) from the atmosphere, representing about 30% of the

global soil C pools with their estimated reservoir of 455 Pg

(1015 g) (Gorham, 1991). This has been achieved by a

relatively small imbalance between production and decay:

only 216% of the net primary production of a peatland

ecosystem gets deposited as peat over centuries or millennia

(Pa iva nen and Vasander, 1994). High water levels and the

consequent anoxia, accompanied with low soil tempera-

tures are considered the major causes for the imbalance. Assuch, the C sink of a peatland is labile, and sensitive to

variations in environmental conditions (Alm et al., 1999a;

Griffis et al., 2000;Bubier et al., 2003a).

Peatlands slowly change along their successional develop-

ment, which is regulated by both allogenic and autogenic

factors (e.g., Klinger and Short, 1996; Hughes and Du-

mayne-Peaty, 2002). Relatively rapid changes in peatland

ecosystem functions may be mediated through land-use

change, and/or climatic warming. In both cases, lowering of

the water level is a key factor (Gitay et al., 2001). Quite

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logically, lowered water levels with the consequent increase in

oxygen availability in the surface soil may be assumed to

result in accelerated rates of organic matter decomposition.

During warm, dry summers that have resulted in temporarily

lowered water levels, substantial C losses from boreal or

subarctic bogs and fens have been observed (e.g., Schreader

et al., 1998; Alm et al., 1999a; Moore et al., 2002). Theseshort-term observations have strengthened the prevailing

paradigm, that persistent lowering of the water level will

reduce the C sink of a peatland, and eventually turn it into a

source of C into the atmosphere (Armentano and Menges,

1986;Silvola, 1986).

In accordance with the paradigm, increased CO2effluxes

have been measured from peatlands where a persistent

lowering of the water level has been induced by ditching

(i.e., those drained for forestry) (Glenn et al., 1993;

Martikainen et al., 1995; Alm et al., 1999b). In great

contrast, an extensive Finnish inventory on long-term (56

decades) changes in drained peatlands indicated that on

average, peat C stores had increased following drainage

(Fig. 1; Minkkinen and Laine, 1998b; Minkkinen et al.,

2002). In individual peatlands, both decreases and in-

creases in peat C stores following long-term drainage have

been observed (Sakovets and Germanova, 1992;Vompers-

ky et al., 1992;Minkkinen and Laine, 1998b;Minkkinen et

al., 1999). In accordance with these contradicting results, in

the few field experiments on drainage-induced changes in

decomposition rates, increase, decrease, and no change

have all been observed (Lieffers, 1988; Minkkinen et al.,

1999;Domisch et al., 2000;Laiho et al., 2004).

Thus, earlier research may be summarized as follows:

after persistent lowering of the water level, a peatland sitemay become a source of C into the atmosphere, remain a

sink, or become a stronger sink. To some extent, these

differences may be linked to peat soil nutrient level

(vegetation type) and climatic conditions (Minkkinen and

Laine, 1998b; Minkkinen et al., 1999); however, the

mechanisms controlling this variation remain unresolved.

The aim of this paper is to synthesize current informa-

tion on the changes, induced by natural or artificial

lowering of the water level, in the biotic and abiotic factors

affecting decomposition processes in peatlands. As the

factors affecting aerobic decomposition in the surface

layers may be considered to experience more changes than

those affecting anaerobic decomposition deeper down,

more emphasis will be given on those. The ultimate goal is

to proceed towards a causal and unifying explanation for

the observed variation in the C sink/source behaviour of

peatlands following lowering of the water level. The review

will limit on changes that might be discernible during the

first 12 centuries, and will not examine peat accumulation

and C sequestration over millennia, which is the time-scale

of peatland succession generally.

Decomposition means mass loss of organic matter as gas

or in solution, caused by either leaching or consumption by

saprotrophic organisms. Terms decomposition, decay,

breakdown and humification have been used in the

literature, some times with varying emphasis (see Clymo,

1984). Here, only the term decomposition is used irrespec-

tive of the original terminology used by the authors in the

studies referred to. Water level refers to the distance from

mire surface to the saturated layer, which shows as a free

water surface in a well, or tube inserted in the peat.

2. Constraints for decomposition in pristine peatlands

Generally, decomposition of organic matter depends on

four factors: substrate quality, environmental conditions,

decomposers present, and nutrient availability. Nutrient

availability is determined by both substrate and environ-

mental characteristics: If decomposer fungi do not get

enough of nutrients from the substrate that they utilize as

an energy source, they may in some cases translocate it

from the surroundings (e.g., Lindahl et al., 2001). These

four factors all interact: environmental conditions and

nutrient availability regulate the vegetation composition

(Wheeler and Proctor, 2000;kland et al., 2001), which in

turn largely determines substrate quality (Hobbie, 1996),

which then, together with environmental conditions and

possibly nutrient availability, regulates the composition of

the decomposer community (Borga et al., 1994; Panikov,

1999).

Substrate quality has usually been described as concen-

trations of C fractions (solubles, holocellulose, lignin), and/

or nutrients. The effects of substrate quality on decom-

position can be studied by incubating different litter types

in constant environmental conditions in the laboratory, or

in similar conditions within a site in the field. Generally,

litters with high concentrations of soluble C compounds,

ARTICLE IN PRESS

Region

1 2 3 4 5

ChangeinsoilC

pool,kgCm-2

-30

-20

-10

0

10

20

30

40Intermediate

Moderately poor

Extremely poor

*

*

*

*

*

*

*

(South) (North)

Fig. 1. Change in total peat soil C pool over 5060 years following

drainage for forestry in pine fens representing different nutrient levels, as

reported by Minkkinen and Laine (1998). Stars show where the change

was significantly different from zero. Region codes: 1

South Finland,2 Central Finland, 3 Eastern Finland, 4 Northern Finland

excluding Lapland, 5 Lapland. Redrawn from Minkkinen and Laine

(1998). Nutrient levels: intermediate fens correspond to mesotrophic,

moderately poor to oligotrophic, and extremely poor to oligo-ombro-

trophic.

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low concentrations of lignin or lignin-like phenolic poly-

mers, and high nutrient concentrations decompose rela-

tively fast (Bartsch and Moore, 1985; Szumigalski and

Bayley, 1996;Preston and Trofymow, 2000;Scheffer et al.,

2001;Thormann et al., 2001;Limpens and Berendse, 2003).

The temporal patterns of these constraints may differ

(Berg, 1984;Sariyildiz and Anderson, 2003), and they mayalso vary among sites (Berg et al., 1993). In the peatland

context, leaf and root litters from herbs and sedges

decompose relatively fast (5175% mass loss during 2

years), moss litters decompose slowly (932% mass loss

during 2 years) and arboreal litters stand in between (mass

loss examples fromSzumigalski and Bayley, 1996;Scheffer

et al., 2001;Thormann et al., 2001). The material used in

the experiments: live, recently dead, or long dead, may

have a considerable effect on the results (Limpens and

Berendse, 2003).

The effects of environmental factors can be studied by

incubating cellulose and/or standard litter materials in

varying conditions. Laboratory studies may yield single

effects of different environmental factors, while field

studies yield the sum of all factors in effect. The most

important environmental conditions that constrain decom-

position in peatlands are generally considered to be oxygen

availability, temperature, and acidity. Accordingly, labora-

tory experiments have shown that within the range of

environmental conditions found in peatlands, CO2 evolu-

tion, an indicator of decomposition, from peat samples

generally increases with lowering water level (Blodau et al.,

2004), increasing pH and increasing temperature (Bergman

et al., 1999).

Extracellular enzymes excreted by certain fungi, actino-myces and bacteria to decompose high molecular weight

complex compounds, are direct indicators of decomposer

activity.Freeman et al. (2001a, 2004) have shown that the

activity of phenol oxidase may be a key regulator of

peatland C cycling. Under low phenol oxidase activity,

phenolic compounds accumulate in the soil, which inhibits

hydrolase enzymes crucial for organic matter decomposi-

tion (Freeman et al., 2001a, 2004). The activity of phenol

oxidase is strongly constrained by the presence of

bimolecular oxygen (Freeman et al., 2001a, 2004). Further,

phenol oxidase activity increases with increasing tempera-

ture (Q10 1.36 in a thermal gradient of 2201C,Freeman

et al., 2001b), and is limited by low pH (Pind et al., 1994;

Williams et al., 2000). Yet,Williams et al. (2000) reported

that phenol oxidase activity in Sphagnum peat, especially,

was regulated less by aeration and more by pH, and when

pH was favourable, phenol oxidase activity depended more

on wetland vegetation type and botanical composition of

the peat than water level. Thus, the role of phenol oxidase

may vary among peatland types.

In general, comparisons among the plentiful field studies

on decomposition rates in different pristine peatland sites

yield few generally valid conclusions on the impacts of the

environmental factors on decomposition. Differences in the

decomposition rates of standard litters have been observed

for sites with differing surface water chemistry: mass losses

have correlated negatively with water levels and, in

contrast to laboratory results, also pH, while positive

correlations with base cation status, soluble N and/or

soluble P have been observed (Verhoeven et al., 1990, 1996;

Verhoeven and Toth, 1995;Szumigalski and Bayley, 1996;

Thormann and Bayley, 1997; Scheffer and Aerts, 2000;Thormann et al., 2001). The patterns have varied among

studies and litter types within studies. This may in part be

due to the decomposer communities being subject to

multiple limitations (Panikov et al. 1997). It would

generally seem that litter type actually exerts a greater

effect on aerobic decomposition than the variation in

environmental conditions among or within peatland sites

(Bartsch and Moore, 1985; Szumigalski and Bayley, 1996;

Scheffer et al., 2001). On the other hand, different litter

types are often produced under different environmental

conditions, and the combined effects of litter quality and

environment on decomposition may be complex (Belyea,

1996). Generally, more readily decomposable litters are

produced under relatively wet conditions (fens vs. bogs,

and hollows vs. hummocks within bogs), and C accumula-

tion may take place because the time for aerobic

decomposition remains short. In contrast, under relatively

dry conditions, less readily decomposable litters are

produced, and the rate of C accumulation is often faster

in spite of the longer time available for aerobic decom-

position (e.g.,Turunen et al., 2002).

There are only few studies that link the decomposers

present in peatland sites to the decomposition of different

substrates. Thormann et al. (2002) observed variation in

the ability of nine species of filamentous microfungiisolated from Sphagnum fuscum1 to decompose different

C compounds or S. fuscum and wood chips. Further,

Thormann et al. (2003)found different fungal assemblages

decomposing different litter types in two peatland sites.

They concluded that litter quality variables were more

important to the variation in the fungal assemblages than

physical or surface water chemistry variables.Williams and

Crawford (1983) have demonstrated high physiological

diversity among peatland microbes: some decomposers

may metabolise well even in cold and very acid conditions.

Coulson and Butterfield (1978) suggested that a major

factor in the slow decomposition ofSphagnummosses and

Eriophorum vaginatum is their unattractiveness to soil

animals. These results again emphasize the role of litter

quality in decomposition processes.

Decomposition rates usually decrease from the surface

downwards, the biggest change taking place when condi-

tions change to anoxic. This is due to both the substrate

becoming more recalcitrant and the conditions becoming

less favourable (Belyea, 1996). There is often a secondary,

or even primary, peak in decomposition rates at the range

of the water level variation (Tuominen, 1981;Santelmann,

ARTICLE IN PRESS

1Nomenclature followsMoore (1982)for vascular plants andKoponen

et al. (1977)for mosses.

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1992; Belyea, 1996), especially if the water level remains

below 20 cm for most of the year, as is often the case in

hummocks (Fig. 2;Tuominen, 1981).

The average water level in northern peatlands generally

varies from a few to around 40 cm below the surface (e.g.,

Gorham and Janssens, 1992;Laine et al., 2004). At a height

growth rate of 2.5mm year1 of the mire surface, it would

take 100 years for litter deposited on the surface to reach a

depth of 25 cm. This calculation is grossly simplified, of

course, as the water level varies seasonally and moss and

root litters are deposited below the mire surface. In any

case, it suggests that in dry sites, many litter types (see

Latter et al., 1998; unfortunately, there are no such data for

mosses) have well reached their maximal mass losses (seeBerg and Meentemeyer, 2002) before entering the anoxic

zone, unlike in wet sites. If lowering in the water level, and

thickening of the oxic zone, were the only changes taking

place, the most important question from the decomposition

point of view would be how much decomposition

potential the substrate exposed for the oxic conditions

would still have.

3. Methods for studying the effects of persistent lowering of

the water level

The effects of persistent lowering of the water level can

be studied using three approaches. Short-term results may

be obtained from laboratory incubations. Because some of

the factors in effect in the nature or their interactions are

always excluded, laboratory incubations fail in producing

reliable long-term results. Long-term results may be

obtained in field studies where the lowering of the water

level has been achieved by ditching. In such studies it is

important to be certain that the ecosystem has not been

manipulated in other ways, like by fertilization, or

vegetation removal or control. Forestry drainage methods

applied in northern Europe and Canada often, but not in

all cases, fulfil this requirement. Mesocosm studies where

relatively large and intact peat monoliths with their natural

vegetation are brought under controlled conditions provide

an intermediate approach. This method allows monitoring

the existing system for a fairly long time; however, it does

not enable plant, possibly also microbial, community

dynamics in the way they would occur in natural

conditions. In the field, invasion of species present in otherhabitats of the peatland or nearby uplands is possible

unlike in the mesocosms. All approaches have their

limitations, and when drawing conclusions one has to

pay special attention to the time scale applied and to the set

of factors that can be considered to show their true

impacts.

Water levels in a peatland usually fluctuate within a

certain range, depending on season and weather condi-

tions. Such fluctuations have their specific effects on

decomposition processes, and these may differ from those

of a persistent change (Moore and Dalva, 1993;Nizovtseva

et al., 1995;Aerts and Ludwig, 1997;Corstanje and Reddy,

2004). The focus in the following will be on a persistent

change in average water levels.

4. Effects on decomposition rates

4.1. Peat soil

There are no studies where in situ dynamics of peat

decomposition following lowering of the water level would

have been reported; however, inferences on soil C balance

have been drawn from soil respiration studies. These are

complicated by the difficulties in separating the specific

sources of the CO2fluxes: autotrophic versus heterotrophicrespiration. Further, what can we consider peat soil?

Possibly a major part of soil respiration originates from the

recently deposited litter instead of peat proper (Malmer

and Holm, 1984). For instance,Hogg (1993)observed that

even under favourable conditions in the laboratory, CO2release from recently deposited Sphagnum litter was nearly

twice as high as that from older material just 10 cm below.

Generally, up to two-fold CO2 emissions have been

observed in sites where water levels have been artificially

lowered, as compared with the controls (Silvola et al.,

1996a;Chimner and Cooper, 2003). Much higher increases

have been observed in laboratory studies (Moore and

Knowles, 1989; Moore and Dalva, 1993; Blodau and

Moore, 2003). A linear relation between CO2emission and

water level has been observed in laboratory conditions

(Moore and Dalva, 1993). In situ, the increase in CO2emission has mostly been seen only with a lowering to a

certain depth, between 10 and 3040cm depending on the

study, with no further increase with a further lowering

(Silvola et al., 1996a;Chimner and Cooper, 2003).Chimner

and Cooper (2003) suggested the lack of easily oxidized

labile C in the deeper soil layers as a reason for this pattern.

This is supported by the results ofHogg et al. (1992)that in

drained samples in vitro the release of CO2 was about 10

times greater from 010 cm peat layer than from 3040 cm

ARTICLE IN PRESS

Water-level

Decomposition potential

Depth

LawnHummock

Surface

Increases

Fig. 2. Schematic presentation of the decomposition potential relative to

water level in a hummock versus a lawn. Based on cellulose decomposition

values presented by Tuominen (1981) for a moderately poor, sedge pine

fen. Note that the actual position of the water level is not the same for the

two cases.

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layer, which they attributed to the relatively large pool of

non-structural carbohydrates in surface samples, deriving

from recently dead plant biomass.

Hogg et al. (1992)suggested that generally, C losses from

peat following water level drawdown essentially depend on

peat quality. Peats that have already been exposed to long

periods of aerobic decomposition may be highly resistantto further decomposition (Bridgham and Richardson,

1992). Consequently, lowered water levels might not cause

a clear increase in C release from peatlands that already

have water tables 20 cm or more below the peat surface for

most of the summer. Many treed bogs and fens would have

such a water regime. In contrast, in peatlands where

organic strata near the surface have been continuously

inundated, the peat with easily oxidizable labile C could

decompose at increased rates. Wet open fens would most

often fall into this category. These postulates are supported

by the results of Silvola et al. (1996a), Minkkinen et al.

(1999)and Lafleur et al. (2005).

It seems evident based on the results reviewed that

lowering of the water level is followed by at least short-

term C losses from the soil, when recently deposited

relatively easily decomposable organic matter is decom-

posed. Further, CO2 stored in the soil profile may be

released as the water-level drops (Moore and Dalva, 1993).

In the laboratory studies ofMoore and Dalva (1993) and

Blodau and Moore (2003) the increased CO2 emissions

lasted for a few weeks following water level drawdown,

when obviously the readily utilizable substrate was

consumed. The Scottish peatlands studied by Hargreaves

et al. (2003) acted as sources of C for 24 years after

draining, ploughing and afforestation; 48 years after treeplanting they acted again as sinks. Nonetheless, there are

little bases yet for estimating the extent and duration of the

increased C losses from different types of peat in situ, that

are caused by different kinds of changes in the hydrological

regime (extent of lowering in the water-level, extent of

fluctuations). Further, the properties of the substrate

available for decomposition will change dynamically with

the inputs of new litter.

Attempts have also been made to directly estimate the

changes that have taken place in soil C pools over time

following drainage (Sakovets and Germanova, 1992;

Vompersky et al., 1992; Minkkinen and Laine, 1998b;

Minkkinen et al., 1999). These comparisons, that involve

many assumptions, cover both the decomposition of the

peat soil that existed at the time of drainage, and the inputs

and decomposition of the litterfall that took place after

drainage. The old peat can only decompose, but

estimating the rate of peat decomposition specifically

would require means for estimating the inputs and

decomposition of the new litters (Vompersky et al.,

1992). Rather surprisingly, the results of these studies

suggest that in most cases, the C balance would have

remained positive. This would mean that the new litter

inputs exceeded the decomposition losses from both the

old peat and the new litters (Sakovets and Germano-

va, 1992; Vompersky et al., 1992). The C pool estimates

include living fine roots, which may contribute up to

1 k g m2 of C (Sjo rs, 1991;Laiho and Fine r, 1996).

Silvola et al. (1996b) estimated that the contribution of

plant roots to the CO2 fluxes was 3545% of total soil

respiration in the middle and late summer at sites with

abundant vegetation. The root contribution correlatedwith tree stand volume. The root proportions were likely

underestimates; the respiration from the plots where roots

had been excluded would still include extra C from the

killed roots during the summer after root isolation. The

lack of a long-term response of CO2 emissions to water

level in the treeless mesocosms ofUpdegraff et al. (2001)

further suggests that tree roots are a major source of CO2emissions in forested sites (see alsoHo gberg et al., 2001). If

the root contribution were about 50% in the sites ofSilvola

et al. (1996a) with the highest soil respiration, the

remainder would be approximately in balance with the

above-ground litter inputs estimated byLaiho et al. (2003)

for a most closely corresponding site (pine-dominated, with

ample moss and shrub vegetation). This comparison lends

some support to the observations mentioned above that in

spite of increased CO2 emissions, the C balance of peat

may not always be negative in drained sites.

4.2. Litter and cellulose

In all studies comparing rates of decomposition in

drained versus pristine peatlands so far, either cellulose

(Lieffers, 1988; Minkkinen et al., 1999) or standard litter

materials (Lieffers, 1988;Domisch et al., 2000;Laiho et al.,

2004) have been used. Thus, the results largely describe thechanges in decomposition potential, i.e., they summarize

the changes in environmental conditions.

Most of these studies have been conducted in treed fens

that have been common objects for forestry drainage. Yet,

the patterns of change that arise from these studies are not

unambiguous. In an Alberta site drained 2 years pre-

viously, both Sphagnum warnstorfiisections and cellulose

decomposed faster than in its pristine counterpart, at the

depth of 30 cm (Lieffers, 1988). At the depth of 10 cm, there

was no difference between sites. In Finland, Minkkinen

et al. (1999) found that cellulose decomposition was in

general faster in sites drained 30 years previously than

in their undrained counterparts, but the depth patterns

of these differences varied among sites (Fig. 3). In the study

of Domisch et al. (2000), drainage increased the mass

loss from Scots pine (Pinus sylvestris) needle litter inCarex-

peat but not in Sphagnum-peat, while mass loss from

Scots pine fine root litter was not clearly affected by

drainage in either peat type. Laiho et al. (2004) found

that mass loss from Scots pine needle litter, as well as

from Scots pine fine root litter incubated at 1020 cm

depth, was faster in a pristine site than in its drained

counterparts, while mass loss from fine roots incubated

at 1020 cm and small roots at either depth did not differ

between sites.

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It is difficult to draw any consistent conclusions from

these studies. Both Lieffers (1988)andLaiho et al. (2004)

attributed the relatively slow decomposition at the surface

of the drained sites to moisture stress. It is noteworthy that

all of these studies dealt with the early phases of

decomposition (max. 1224 months) only. Long-term

patterns may differ from short-term observations, and willbe more crucial from the C sequestration point of view

(Latter et al., 1998; Moore et al., 2005; Prescott, 2005).

Further, with the exception of the work by Laiho et al.

(2004), the experimental designs did not include replica-

tions over time, i.e., the impacts of annual variation in

climatic factors could not be accounted for.

5. Effects on the abiotic and biotic factors regulating

decomposition processes

5.1. Physical and chemical environment for decomposition

Following a lowering of the water level, the water

content in the surface peat will decrease. Simultaneously,

the air-filled porosity and oxygen content increase (Boggie,

1977), and the oxic limit extends deeper (La hde, 1969;

Silins and Rothwell, 1999). These changes will generally

improve the conditions for the aerobic decomposition of

organic matter, which is faster than anaerobic decomposi-

tion (e.g.,Bergman et al., 1999;S antruckova et al., 2004).

Simultaneously, the density of the surface peat may

increase due to several factors (Minkkinen and Laine,

1998a;Silins and Rothwell, 1998;Laiho et al., 1999). Peat

compaction, in turn, is associated with changes in pore size

distribution: the proportion of small pores with high water

retention capacity increases (Pa iva nen, 1973; Silins and

Rothwell, 1998). This may partly counteract the changes in

water and oxygen contents initially brought about by a

lowering of the water level (Rothwell et al., 1996).

If the water level drops deep enough, moisture stress may

counter the positive effects of drainage on decompositionin the topmost soil layers (Lieffers, 1988). Peat soils

generally have a capillary fringe that reaches the surface

when the water level is within 3040 cm, only in highly

decomposed soils may it reach the surface when the water

level is at 60 cm (Verry, 1997).Laiho et al. (2004)showed

low soil matric potentials in the top 5 cm of soil in a

drained fen site when the water level was 40 cm below the

surface. Even if the topmost soil becomes sub-optimally

dry, it seems likely that there is a layer between the dry

surface and the new water level where increased aeration

and decreased moisture result in improved conditions for

aerobic decomposition.

There are also other noteworthy changes in the abiotic

conditions that may affect decomposition. First of all, soil

thermal properties and temperature regime change. Initi-

ally, during 13 years, the growing-season temperatures in

surface peat may increase (Lieffers and Rothwell, 1987;

Lieffers, 1988); however, along with an increase in tree

cover, temperatures in drained boreal peatlands remain

colder than in pristine sites (Ho kka et al., 1997;Vena la inen

et al., 1999). Minkkinen et al. (1999) observed that this

difference extended at least to a depth of 50 cm, at which

depth the temperature remained below +51C approxi-

mately a month longer in the drained site than its pristine

counterpart. The temperature sum (0 1C threshold) of thesurface soil (025 cm) may be 400d.d. higher in pristine,

wet conditions than after long-term water-level drawdown

(Laine et al., 2004) (Fig. 4). Such a difference in air

temperature sums would correspond to a 500700 km

geographical distance in northsouth direction. Also, soil

frost tends to melt later in drained than in undrained sites

(Eurola, 1975). These indirect impacts of lowered water

levels are noteworthy as they may counteract the effects of

increasing air temperatures.

Further, the acidity of the surface soil increases

following water-level drawdown (Lukkala, 1929; Laiho

and Laine, 1990; Laine et al., 1995a; Minkkinen et al.,

1999). This has been attributed to several factors such as

reduction in the inflowing groundwater that has a

neutralizing effect in pristine minerotrophic sites, enhanced

oxidation of both organic and inorganic compounds, and

increased uptake of base cations by tree stands (Laine et

al., 1995b). In the treed fens studied byLaiho and Laine

(1990)the decrease in soil pH extended down at least to the

2535 cm layer, being in the range of 0.51 units inCarex

peat during 5 decades, and less in Sphagnumpeat (dry peat

to 0.01 M CaCl2 1:2.5).

The changes in temperature and soil acidity may be

considered to have a retarding effect on decomposition

(e.g., Ivarson, 1977; Bergman et al., 1999), even though

ARTICLE IN PRESS

Decomposition potential

Depth

Surface

Increases

Before

After, option 1

After, option 2

Old water-level

New water-level

Fig. 3. Schematic presentation of the change in the decomposition

potential caused by a lowering of the water level. Based on cellulose

decomposition values presented by Minkkinen et al. (1999), and needle

decomposition values presented by Laiho et al. (2004), for moderately

poor, sedge pine fens. The Before case relates to old water-level, and

the After-cases, which represent situations where the surface layer either

does (option 1) or does not (option 2) experience moisture stress or other

changes retarding decomposition (hummock conditions), relate to new

water-level.

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many peatland microbes may tolerate quite acid conditions

(Williams and Crawford, 1983). However, environmental

factors have interactions that complicate the interpreta-

tions of these changes. For instance, the effect of changes in

temperature depends on moisture conditions. Hogg et al.

(1992) observed no change in soil respiration with higher

incubation temperature under inundated conditions. On

the other hand, Laiho et al. (2004) observed that higher

temperature sums during fixed-length incubation periods

led to lower mass losses from needle and fine root litter in

drained treed fens. They concluded that this was because

higher temperatures enhanced moisture stress in the surface

layers of drained sites. This conclusion was supported by

positive correlations between mass losses and summertime

precipitation sum (Laiho et al., 2004).

The interactions between temperature and moisture

should also be considered when estimating the effects of

increasing temperatures, caused by climate change, on

decomposition rates. An increase in temperature will have

an unambiguously positive effect on decomposition only

when the moisture conditions are favourable. In sub-

optimally dry conditions, decomposition may be retarded

because of moisture stress, and in waterlogged conditions

because of lack of oxygen (La hde, 1969; La hde, 1971;

Hogg et al., 1992).

Interestingly, and surprisingly, Coulson and Butterfield

(1978) and Moore et al. (2005) have reported that there

were no consistent differences in mass loss rates of several

standard litter materials between upland and nearby

peatland sites. This emphasizes again the significance of

litter quality in addition to environmental conditions for

the progress of decomposition.

5.2. Vegetation and organic matter inputs

There are two major biotic agents in the C cycle of a

peatland ecosystem: vegetation, which produces the or-

ganic matter (litter), and decomposers, which consume the

organic matter to a varying extent. Vegetation composition

is largely determined by water level and the solution pH

(Wheeler and Proctor, 2000; kland et al., 2001). A

persistent change in the water level induces acclimatization

and adaptation, first within the existing community

(Weltzin et al., 2000, 2003), followed by slower but more

drastic changes when species better adapted to the new

conditions gain dominance (Laine et al., 1995a). Changes

in both species abundances and species composition may

be critical turning points for the C balance of a site

(Malmer et al., 2003). Litter quality is the key: different

ARTICLE IN PRESS

Fig. 4. Peat temperature (bottom) and temperature sum (accumulated daily average temperatures40 1C) (top) at different depths in a pristine, moderately

poor, sedge fen (left) and its counterpart drained for forestry 30 years earlier (right). The undrained site was treeless, while the drained one had developed a

tree stand of about 110 m3 ha1 following drainage. Source:Laine et al. (2004; see also Minkkinen et al., 1999).

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litter materials may have vastly differing rates of decom-

position (Hobbie, 1996;Thormann et al., 2001).

Generally, following a lowering of the water level species

adapted to wet conditions rapidly decline, while hummock

species benefit. Sedges such as Carex lasiocarpa and

C. rostrata, which are important for peat formation in

many fens, are the first to suffer from drought (Bubier etal., 2003a,b), and disappear relatively fast after a persistent

drop in the water level (Laine et al., 1995a). Eriophorum

vaginatumis generally the only graminoid that can thrive in

dry sites, until the shading from the tree stand becomes too

much at canopy closure. Sphagnummosses, another species

group important for peat formation, generally decline

following a persistent lowering of the water level, while

forest mosses such as Pleurozium schreberi and Hyloco-

mium splendensincrease in abundance (Laine et al., 1995a).

In some treed fen types, it has been observed that both the

coverage (Laine et al., 1995a) and biomass (Laiho et al.,

2003) of Sphagnum mosses may remain high, thanks to

hummock species such asS. angustifolium,S. magellanicum

and, especially, S. russowii(Laine et al., 1995a).

Shrubs are the first to benefit from lowered water levels

(Laine et al., 1995a; Bubier et al., 2003b; Laiho et al.,

2003), while trees may gain dominance over a longer period

of time. Following persistent lowering of the water level,

mire species are at least to some extent gradually replaced

by forest species (Laine et al., 1995a;Vasander et al., 1997).

The extent and rapidity of this secondary succession

depend on the nutrient regime and initial wetness of the

site. The changes are the greatest in initially wet, nutrient-

rich sites (Laine et al., 1995a). In extremely nutrient-poor

ombrotrophic conditions the changes are small: in initiallydry sites they may remain almost non-existent (Vasander,

1982), while in initially wet sites or microforms, lichens

may replace the specialized mosses (Jauhiainen et al.,

2002).

Total plant biomass in peatland sites has a strong

positive correlation with tree stand biomass (or volume),

while the biomass of lower vegetation layers is negatively

correlated with tree stand biomass (Reinikainen et al.,

1984). Thus, total plant biomass generally increases

following lowering of the water level (Reinikainen et al.,

1984; Laiho et al., 2003) unless the site is too poor in

nutrients to support tree growth (Vasander, 1982). Total

production is also higher in dry than wet sites, again with

the exception of the most nutrient-poor sites (Reinikainen

et al., 1984).

The effects of vegetation changes on the quantity and

quality of litter inputs has received little attention, even

though it was recognized already byLieffers (1988)that the

type and mass of litter will change following drainage,

which should have an effect on decomposition processes.

Laiho et al. (2003) estimated that in treed fens, the total

annual litterfall (including moss and below-ground litter)

did not change radically during 60 years following

drainage, and ranged from ca. 600 g m2 year1 (2025

years after drainage) to 960 g m2

year1

(8 years after

drainage) (Fig. 5). In contrast, there were dramatic changes

in litterfall composition. In pristine state, half of the total

annual litterfall derived from sedges, one third from

mosses, and less than 20% from trees and shrubs. Within

20 years after water-level drawdown, shrubs and trees had

become the major litter source, producing more than 80%

of the total litterfall. At 55 years after drainage, 68% of

total litterfall derived from trees and shrubs, and 30% from

mosses. There was also a change in the composition of the

arboreal litterfall in the course of time, as the proportions

of woody debris and cones increased.

The changes in the tissue type composition of the

litterfall are likely to have a negative effect on decomposi-

tion rates. Generally, arboreal litters decompose slower

than graminoid and herbaceous litters (Taylor et al., 1991;

Hobbie, 1996;Moore et al., 1999;Thormann et al., 2001).

This holds for foliage also, but woody materials: branches,

cones and eventually tree trunks, decompose especially

slowly (Taylor et al., 1991; Hobbie, 1996; Laiho and

Prescott, 1999).

Of Sphagnum mosses, the species adapted to dry

conditions (e.g., S. fuscum, S. magellanicum) decompose

slower than the species of wet habitats (e.g., S. cuspidatum,

ARTICLE IN PRESS

Fig. 5. Evolution of litter inputs into a moderately poor, sedge pine fen

following persistent lowering of the water level. Based on a chronose-

quence study. Source: Laiho et al. (2003). FWD fine woody debris,

other nonw.

other nonwoody debris.

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S. fallax) (Johnson and Damman, 1991; Limpens and

Berendse, 2003). Sphagnum mosses are generally consid-

ered as poor, slowly decomposing materials (Standen and

Latter, 1977), but also the forest mosses common in

drained sites (e.g., Pleurozium schreberi, Hylocomium

splendens) have proven to decompose very slowly (Hobbie,

1996; Karsisto et al., 1996). Thus, the abundance andvigour of the moss layer may significantly affect the carbon

balance of a given site.

While in pristine peatlands sedge roots may grow down

to depths of more than 2 m (Saarinen, 1996), the arboreal

communities dominating dry/drained sites are generally

shallow-rooted (Paavilainen, 1966a,b). This means that

following water level drawdown, there may be a reduction

in the priming effect of root exudates and fresh root litter

on the anaerobic decomposition in deeper peat layers.

5.3. Decomposer communities

The structure of the microbial and faunal communities

varies with the peatland plant community (Borga et al.,

1994;Drouk, 1995;Fisk et al., 2003). Thus, it is likely that

such changes in hydrology that affect the plant community

will also induce adaptation in the microbial and faunal

communities (Markkula, 1986a,b), and the patterns of

adaptation may change over time since the hydrological

change.

Unfortunately, only sporadic information exists on the

changes in the decomposer communities and their func-

tioning caused by changes in the water level. Paarlahti and

Vartiovaara (1958) reported an increase in cellulose-

decomposing fungi and bacteria down to a depth of about30 cm in intensively drained sites.Karsisto (1979)observed

an increased number of aerobic bacteria in sites where the

water level had been lowered down to 70 cm, as compared

to similar sites with water levels in 10 and 30 cm depths.

The increase was limited to the 010 cm peat layer. In

contrast, she observed generally lower lengths of fungal

mycelia in the sites with the deepest water levels. As yet, the

responses of decomposer communities to changing water

levels have not been analysed using PLFA profiles

(phospholipid fatty acids; Ratledge and Wilkinson, 1988)

or molecular methods.

The food webs in peatlands are, as a whole, still

relatively poorly known. It has been suggested that

enchytraeids drive the processes of early stage decomposi-

tion in peat soils (Standen, 1973, 1978;Cole et al., 2000).

The majority of soil invertebrates, including enchytraeids,

occur in the surface horizons where they are vulnerable to

changes in environmental conditions (Briones et al., 1997;

Cole et al., 2002a). They thrive in a relatively moist and

cool environment (Briones et al., 1997), and are vulnerable

to drought (e.g., Cole et al., 2002a). During dry, warm

periods, some species such as Cognettia sphagnetorummay

migrate downwards (Springett et al., 1970; Briones et al.,

1997); however, they may not tolerate extended periods of

anoxia (Healy, 1987). Further, the more decomposed

substrate of deeper soil layers may not be suitable for

enchytraeids. Briones and Ineson (2002) showed that

enchytraeids assimilated carbon that had been removed

from the atmosphere 510 years earlier. Correspondingly,

Cole et al. (2000) suggested that downward migration of

enchytraeids during drought might have little effect on

decomposition.The responses of soil mesofauna to potential climate

change have been studied in the blanket peats in UK

(Briones et al., 1997, 2004; Cole et al., 2002a). Only

temperature changes have been considered; no water level

treatments have been applied.Cole et al. (2002b)suggested

that soil warming might reduce the functional role of

enchytraeids with respect to their ability to enhance

decomposition.

Silvan et al. (2000) observed that the total numbers of

enchytraeids, collembolans and mites correlated positively

with water level depths in a drainage-succession continuum

of sedge pine fens. In sites where the water level was at a

depth of 2530 cm, the populations were generally about

ten times higher than in sites where the water level was at

5 cm. Simultaneously, the proportion of oribatid mites

decreased and that of collembolans increased.Laiho et al.

(2001)further observed that when the average water level

of a site was below 20 cm, it did not have a significant effect

on the within-site distribution of mites, but still affected

that of enchytraeids and, especially, collembolans. Species

composition, activity, or role in the food web, were not

accounted for in these papers.

The implications of the observed changes in decomposer

communities following lowering of the water level for

organic matter decomposition are currently still somewhatdifficult to estimate. It seems likely that they would be

enhancing decomposition.

6. Synthesis from a C sink/source point of view

Based on the results reviewed, it is clear that although

high water levels in peatlands (relative to upland sites)

result in C sequestration, within a peatland, the relations

between water level, decomposition and C sequestration

are not that simple (see alsoWaddington and Roulet, 2000;

Bubier et al., 2003b).

Mechanistically, whether a peatland will remain a C

sink, become a stronger sink, or become a source of C to

the atmosphere will depend on 1) the rate of decomposition

of the old, pre-water-level-lowering, peat, and 2) the

rates of inputs and decomposition of the new organic

matter entering the system as litter produced under the new

environmental conditions (Fig. 6). Quite simply, if the

accumulation (inputs decomposition losses) of the new

organic matter exceeds the decomposition losses from the

old peat, the peatland will remain a sink of C. If not, then

the peatland will become a source of C to the atmosphere.

The old peat can only decompose. The C balance of the

new litter inputs will largely depend on the vegetation

composition. Easily decomposable tissue types may be

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preferred by decomposers over the older substrate (Wil-

liams and Crawford, 1983), while in case of lowered litter

quality (e.g., tree litter vs. graminoid litter), the situation

may be the opposite. The role of the changing vegetation in

the C cycling following a persistent lowering of the water

level has in previous research been largely neglected.

The first postulate regarding the effects of lowered water

levels on decomposition processes and C sink/sourcebehaviour is that long-term changes differ from short-term

changes that merely reflect a disturbance in the old system

adapted to the pre-water-level-lowering conditions. In a

short term, the disturbed system will probably always lose

C. An essential question here is, for how long? Obviously,

the C loss from peat is at its greatest during and

immediately after the lowering in the water level, when

drought is not yet limiting decomposition, and slows down

with time (e.g., Blodau and Moore, 2003), due to

decreasing substrate quality. The photosynthetic flow of

C into the soil, in turn, may also decrease (Bubier et al.,

2003a,b; Blodau et al., 2004), but may recover with the

response of the plant community, such as transition from

sedges to shrubs that may occur rather rapidly (Bubier et

al., 2003b). The rapidity of the response of the plant

community may depend on the site nutrient regime

(Tuittila, E.-S., Alm, J. and Laine, J. unpublished data).

In Scotland, newly drained peatlands acted as sources of C

during the first 4 years, after which they became sinks again

(Hargreaves et al., 2003).

The second postulate is that long-term changes in the C

cycle result from several adaptive mechanisms of the

ecosystem to the new hydrological regime. Hence, the

long-term changes may inherently vary among peatland

types, climates, and extents of change in the water level.

The determinants must in one way or the other derive from

the interactive effects of the changes in not only water level

and aeration (Blodau and Moore, 2003), but also vegeta-

tion composition and organic matter inputs (Bubier et al.,

2003b; Laiho et al., 2003), and possibly soil temperature

and acidity (Minkkinen et al., 1999).

The reviewed studies emphasize the paramount impor-tance of organic matter quality in determining the rate of

decomposition in peatlands (see alsoBauer, 2004). Actually,

such observations led alreadyCoulson and Butterfield (1978)

to suggest that the position of the water level or the pH

differences in soils were of only indirect consequence for the

rates of decomposition, through determining the composi-

tion of the plant community. However, when evaluating the

effects of persistent lowering of the water level on C cycling,

accounting for changes in species composition only would be

an oversimplification just like assuming that the change in

water level is the only determinant. On deep-peat sites the

lower parts of the peat deposit remain oxygen-deficient

despite of a drop in the water level, and maintain a zone

where new material may be accumulating or at least old

material may be preserved. Thin-peated sites, if they lose

altogether an anoxic layer, may be at the greatest risk for

losing considerable amounts of C.

The few studies that have so far systematically looked at

the effect of water level on decomposition within one

vegetation type, were all done on rather nutrient poor treed

fen or bog sites in southern or middle boreal conditions

(Lieffers, 1988; Minkkinen et al., 1999; Domisch et al.,

2000; Laiho et al., 2004). From these studies it may be

concluded that in such peatland types, the site may turn

into a large hummock-system where several factors,including litter quality, relative moisture deficiency, lower

soil temperature, higher acidity, and in deeper layers also

oxygen deficiency, may interact to constrain organic matter

decomposition. This may explain the continuing sink

behaviour that has been observed in these kinds of sites.

Currently, we do not know what are the critical limits in

the properties of vegetation (litter inputs) and peat that

switch the site either to a source or a sink. A critical factor

may be the quality and thickness of the intermediate

peat layer where the rate of decomposition is increased in

any case (Fig. 3).

7. Conclusions

Based on the literature reviewed, the following critical

gaps have been identified in our knowledge of the C cycle in

peatlands under change. There is a lack of information on:

1. how the amounts and quality parameters of litter inputs

change in different peatland sites after short- and long-

term change in the water level,

2. how the litters produced by the successional vegetation

communities decompose under the changed environ-

mental conditions following persistent lowering of the

water level in long-term,

ARTICLE IN PRESS

Before After 1 After 2 After 3 After 4 After 5

SoilC

pool,%

0

20

40

60

80

100

120

Peat

Litter

Fig. 6. Schematic presentation of the sink/source behaviour of a peatland

site following a persistent lowering of the water level. Peat C pool in the

peat deposit; Litter C balance (input-decomposition) of annual litter

inputs. Before before the lowering of the water level. The different

After-options depict situations where peat decomposition clearly increases

(1, 2 and 5) or is little affected (3 and 4), and the C balance of litter inputs

remains the same (1 and 3), increases (2 and 4) or decreases (5). The soil Cpool (y-axis scale) is in per cent of the pool in peat before water-level

drawdown. Thus, an After-column below the dotted 100% line implies a

switch to a C source, while a column above this line implies continued sink

function.

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3. how the changes in litter inputs affect the community

structure and functioning of microbes responsible for

the aerobic decomposition of organic matter, and

4. how the quality parameters of peat affect its decom-

position rates under changed environmental conditions.

Further, the interactions of different environmentalfactors have not been systematically analysed: how are

different water level regimes related to oxygen supply,

temperature, acidity, and nutrient supply within a given

site? The variations in physical factors could be simulated

with existing models (Granberg et al., 1999; Zhang et al.,

2002) but we need test data for situations with lowered

water levels.

Consequently, we need more successional studies where

we consider the interactions of all changing factors

(vegetation, microbes, several abiotic) on sites with

different old peat substrates. Especially, we need

experimental research that has been designed to contribute

to an overall synthesis.

Each case study will bring in valuable new information

on changes taking place in single peatland types and

climates. Yet, because of the multiple interactions between

the factors in effect, generalizations should not be based on

limited data sets. Moreover, we should generally be more

specific when discussing decomposition of organic matter,

and attribute our statement to both the subject of

decomposition (peat, cellulose strips, litter from different

plant species) and the time scale involved.

Acknowledgements

My sincere thanks are due to Hannu Fritze, Krista

Jaatinen, Marjut Karsisto, Veikko Kitunen, Jukka Laine,

Kari Minkkinen, Taina Pennanen and Petra Va vrova for

inspirational discussions, Tim Moore, Timo Penttila , Eeva-

Stiina Tuittila and Harri Vasander for critical and

constructive comments, and the Academy of Finland for

funding (projects 203585, 205090).

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