Controls of bioavailability and biodegradability of dissolved organic matter in soils

Controls of bioavailability and biodegradability of
dissolved organic matter in soils
Bernd Marschner
a,
*
, Karsten Kalbitz
b
a
Department Soil Science/Soil Ecology, Ruhr-University Bochum, D-44780 Bochum, Germany
b
Department of Soil Ecology, Bayreuth Institute for Terrestrial Ecosystem Research (BITO
¨
K),
University of Bayreuth, D-95440 Bayreuth, Germany
Received 11 March 2002; accepted 9 December 2002
Abstract
In soils, dissolved organic matter (DOM) is probably the most bioavailable fraction of soil
organic matter, since all microbial uptake mechanisms require a water environment.
Bioavailability describes the potential of microorganisms to interact with DOM. It is a
prerequisite for biodegradation and can be restricted, if DOM is present in small pores or within
soil aggregates and therefore not accessible for microorganisms. DOM biodegradation is defined
as the utilisation of organic compounds by soil microorganisms quantified by the disappearance
of DOM or O
2
or by the evolution of CO
2
. The controlling factors for DOM biodegradability
can be divided into three groups, namely, intrinsic DOM quality parameters, soil and solution
parameters and external factors. DOM characteristics that generally enhance its biodegradability
are high contents of carbohydrates, organic acids and proteins for which the hydrophilic neutral
fraction seems to be a good estimate. In contrast, aromatic and hydrophobic structures that can

also be assessed by UV absorbance decrease DOM biodegradability, either due to their
recalcitrance or due to inhibiting effects on enzyme activity. Effects of solution parameters such
as Al, Fe, Ca and heavy metal concentrations on DOM biodegradability have been documented
in various studies, however with different, sometimes conflicting results. Inhibitory effects of
metals are generally attributed to toxicity of the organic complexes or the free metal ions. In
contrast, the enhanced degradability observed in the presence of metal ions may be due to
flocculation, as larger structures will provide better attachment for microbial colonies. As
degradation is dependent on microbial activity, the composition and density of the microbial
population used in the degradation studies also influence biodegradation. Site-specific factors,
such as vegetation, land use and seasonality of meteorological parameters control DOM
composition and soil and soil solution properties and therefore also affect its biodegradability.
0016-7061/02/$ - see front matter D 2002 Elsevier Science B.V. All rights reserved.
doi:10.1016/S0016-7061(02)00362-2
* Corresponding author. Tel.: +49-234-3222108; fax: +49-234-3214469.
E-mail address: (B. Marschner).
www.elsevier.com/locate/geoderma
Geoderma 113 (2003) 211 – 235
The major obstacle for a better understanding of the controls of DOM biodegradability is the
lack of a standardised methodology or at least systematic comparisons between the large number
of methods used to assess DOM biodegradability.
D 2002 Elsevier Science B.V. All rights reserved.
Keywords: Bioavailability; Biodegradation; Dissolved organic matter; DOC
1. Introduction
In the past 10 years, much progress has been made in the understanding of dissolved
organic matter (DOM) functions and dynamics in soils. Today, it is commonly acknowl-
edged that DOM can enhance the solubility and mobility of metals and organic
compounds (Blaser, 1994; Piccolo, 1994; Zsolnay, 1996; Marschner, 1999) and thus
contributes to pollutant transport or to micronutrient availability. In the presence of DOM,
weathering rates can be accelerated (Raulund-Rasmussen et al., 1998), and DOM plays a
central role during podsolisation (Lundstro¨m et al., 1995). Furthermore, DOM contains

organically bound nutrients such as N, P and S, and DOM dynamics will therefore also
affect their mobility and availability (Kalbitz et al., 2000; Kaiser et al., 2001a).
DOM is also a substrate for microorganisms. In soils, DOM may be the most important
C source since soil m icroorganisms are basically aquatic and all microbial uptake
mechanisms require a water environment (Metting, 1993). Furthermore, the soluble state
is presumably a prerequisite for the diffusion of substrates through microbial cell
membranes so that the degradation of solid phase organic matter or large molecules can
only occur after dissolution or hydrolysis by exoenzymes. The initial phase of litter
decomposition is also strongly related to the amount of soluble compounds in the litter
(Williams and Gray, 1974). This was also shown by Marschner and Noble (2000), where
CO
2
release from a soil supplemented with different plant litters could largely be explained
by the disap pearance of DOC (Fig. 1). Similar results were obtained with soils incubated at
different temperatures (Marschner and Bredow, 2002). Cook and Allen (1992) also report
positive relationships between initial DOC concentrations and CO
2
release during the first
5 weeks of a long-term incubation experiment. However, at later stages, this relationship
no longer existed which they attributed to the depletion of degradable DOM compounds.
Several other authors have found close correlations between DOM concentrations and
denitrification potentials or rates (Bijay-Singh et al., 1988; Isermann and Henjes, 1990; Pu
et al., 1999), thus indicating that the availability of biodegradable DOM may be a
prerequisite for creating reducing condition in soils or in certain soil compartments
(Zsolnay, 1996). On the other hand, Kalbitz et al. (2003) found no evidence that DOM
extracted from Oa, and A horizons is the most biodegradable fraction of soil organic
matter.
DOM degradation is also an important process controlling DOM dynamics in soils.
DOM inputs into the mineral soil generally greatly exceed DOM outputs with seepage .
Until recent ly, this was mainly attributed to DOM retention through sorption (Guggen-

berger et al., 1998). However, some newer calculations indicate that total C pools should
then be several orders o f magnitude higher than generally observed in the field
B. Marschner, K. Kalbitz / Geoderma 113 (2003) 211–235212
(Guggenberger and Kaiser, 2003; Moore, 1997). Therefore, the mineralisation rate of
DOM in subsoils must be much higher than thought previously (Qualls and Haines, 1992;
Guggenberger et al., 1998), and thus, the biodegradation of DOM or of former DOM
sorbed to mineral surfaces is the most likely explanation for the generally low DOM fluxes
towards the groundwater reported by Michalzik et al. (2001).
In their extensive review on DOM dynamics in soils, Kalbitz et al. (2000) also point out
that the mechanisms and controls of DOM degradation in soils are still poorly understood.
If the earlier stated assumption is correct, that the soluble state is a prerequisite for the
uptake and degradation of organic matter by microorganisms, then DOM should play a
key role in the stabilisation and destabilisation of soil organic matter, and thus, in C
dynamics and C pools of soils. Sollins et al. (1996) present a conceptual model of SOM
stabilisation and desta bilisation for which they differentiate between three general sets of
characteristics affecting the stability of organic matter: recalcitrance, interactions and
accessibility. For the co nceptual understanding of mechanisms and controls, this approach
is very helpful. However, one has to bear in mind that often, several mechanisms and
processes interact to determine the stability or biodegradability of organic mat ter in soils.
In this paper, factors and mechanisms are reviewed that control the microbial
degradation of DOM in soils. The term ‘‘DOM’’ will be used for all organic substances
smaller than 0.45 Am that are suspended in aqueous solutions. Strictly speaking, the term
DOM can only be applied to organic matter in soil solutions extracted with lysimeters. In
most studies where soluble organic matter is obtained after extraction from the soil with
batch or percolation methods, this is not the truly dissolved phase, but the potentially
soluble. Zsolnay (1996) therefore suggests to clarify this by using the term ‘‘WEOM’’
(water-extractable organic matter). Since this review deals with the degradability of
organic substances in the solution phase, a differentiation between the various methods
whereby this solution phase was obtained is not regarded as essential for the problem.

Fig. 1. Relationship between the change in water-extractable soluble organic compounds (DOC) and cumulative
CO
2
evolution in an Australian pasture topsoil during a 21-day incubation with different plant litter materials
(Marschner and Noble, 2000).
B. Marschner, K. Kalbitz / Geoderma 113 (2003) 211–235 213
2. Bioavailability versus biodegradability
In pharmaceutical and toxi cological studies with mammals, the term ‘‘bioavailability’’
is used to characterise the amount of a substance ingested and retai ned in the organism and
thus becomes available for metabolic use. It is therefore not a measure for the actual
utilisation of this substance. For organic molecules such as DOM compounds, this means
that their uptake, i.e., bioavailability, must not necessarily result in their breakdown to
smaller entities or to complete mine ralisation. On the other hand, microorganisms excrete
exoenzymes that promote extracellular degradation of compounds that are otherwise not
bioavailable according to the above definition. Therefore, in the context of DOM, the term
bioavailability describes the potent ial of microorganisms to interact with these substances.
As a measure for the actual utilisation of organic compounds by soil microorganisms,
the term ‘‘biodegradability’’ is chosen. In a strict sense, this still encompasses two
alternative or sequential processes:
(1) microbial uptake or breakdown of the original compounds which are then used for the
biosynthesis of microbial cell materials.
(2) complete mineralisation to obtain energy and inorganic nutrients.
Depending on the analytical tools used to moni tor the degradation process, these two
processes a re considered to different degrees. If microbial utilization of DOM is
determined by the increase in microbial biomass, then only the assimilated organic carbon
(AOC) is considered (Escobar and Randall, 2001), while the min eralized fraction is
neglected. If DOC disappearance is used as a measure for biodegradation, it is not possible
to differentiate between microbial incorporation and mineralisation.

Another aspect that points to the complexity of this issue is illustrated in the study by
Amon and Benner (1996) where low-molecular DOM (< 1000 Da) was less degradable
than high-molecular DOM. However, bacterial growth efficiency was much higher with
the low-molecular DOM fraction, thus indicating that this seemingly less-degradable
fraction contained more compounds needed for bacterial biomass production.
AOC is a measure for the ability of water to support heterotrophic growth (Escobar and
Randall, 2001). Therefore, AOC is mainly an important parameter for waterworks. It
represents only a small portion of the entire biodeg radable DOM and will not be further
discussed in this paper mainly dealing with soil DOM. We will focus on biodegradability
of terrestrial DOM, i.e., the utilisation of organic compounds by soil microorganisms
quantified by the disappearance of DOM or O
2
or by the evolution of CO
2
.
3. Methods for the determination of DOM biodegradability
As indicated in the previous section, no generally accepted standard method for the
determination of DOM biodegradability exists so far. A method published by the
International Organization for Standardization determines the biochemical oxygen demand
(BOD) of aqueous media, which is based on a 5-day incubation, using solid sewage sludge
as an inoculum (ISO 10707, 1994). However, this method is mainly used to determine the
B. Marschner, K. Kalbitz / Geoderma 113 (2003) 211–235214
BOD in effluents from sewage treatment plants that have to meet certain BOD levels in
many countries.
The methods used by soil scientists, groundwater hydrologists, limnologists or ocean-
ographers to quantify the biodegradable organic carbon are quite diverse (Table 1). Most
studies are conducted in cultures where any of the following parameters may vary:
– type of incubation (batch culture, flow-through bioreactor)
– type and size of incubation vessel
– shaking

– duration of incubation
– initial DOC concentration
– nutrient additions
– type and amount of inoculum
– temperature
– measure for biodegradation (CO
2
efflux, DDOC or DTOC) and frequency of
measurements
In addition, data analysis and documentation will either identify different pools of
biodegradable DOM (labile, semi-labile, stable, fast, slow) based on degradation rates, or
simply quantify the amounts of DOC mineralised or remaining after a certain time period.
As a matter of fact, in this review, no two studies performed in different laboratories used
the same set of parameters for the determination of DOM biodegradability in their batch
experiments. This means that the reported results cannot be compa red with each other,
which is a major obstacle for scientific discussions and progress.
Of all the parameters listed above, two seem to be most crucial for the quantification of
DOM biodegradability: duration of incubation and measure for biodegradation.
Many long-term incubations (>10 days) showed that DOM generally consists at least of
a rapidly degradable fraction (fast BDOM or labile DOM), a fraction that is degraded more
slowly, and the recalcitrant fraction that remains in solution even after very long
incubation periods (up to 180 days). Little is known about the nature of the compounds
in these different DOM pools, but it is general ly assumed that the labile DOM consists
mainly of simple carbohydrate monomers (i.e., glucose, fructose), low-molecular organic
acids (i.e., citric, oxalic, succinic acid), amino acids, amino sugars and low-molecular-
weight proteins (Lynch, 1982; Qualls and Haines, 1992; Guggenberger et al., 1994; Ku
¨
sel
and Drake, 1999; Kaiser et al., 2001b; Koivula and Ha¨nninen, 2001). These compounds
can directly be utilised by a large number of different organisms and therefore do not

require a special set of enzymes (Lynch, 1982).
The slowly degradable or relatively stable DOM fraction probably contains polysac-
charides (i.e., breakdown products of cellulose, hemicellulose) and other plant or micro-
bially derived compounds or degradation products that require special tools for
degradation. These enzymes are probably only produced when labile substrates are no
longer available or they are limited to few organisms that therefore do not need to compete
for the labile substrates (k- vs. R-strategists according to Paul and Clark, 1996). Even the
so-called recalcitrant DOM fraction is not fully nondegradable as it woul d otherwise
accumulate in subsoils to a much higher degree than observed (Moore, 1997). However,
B. Marschner, K. Kalbitz / Geoderma 113 (2003) 211–235 215
Table 1
Summary of methods used in different studies for the determination of the biodegradability of DOM or DOM fractions
Type of incubation Biodegradation Additives Duration of incubation
measure
V 1 day
1 – 7 days 1 – 4 weeks H4 weeks
Batch incubation in
solution culture
DOC or TOC
a
inoculum Pinney et al., 2000
b
Benoit et al., 1968;
Block et al., 1992
c
;
Nelson et al., 1994;
Boyer and Groffman,
1996; Volk et al.,
2000

b
; Escobar et al.,
2001
b
Zsolnay and Steindl, 1991
d
;
Qualls and Haines, 1992;
Boissier and Fontvieille, 1993;
Raymond and Bauer, 2001
inoculum +
nutrients
Zsolnay, 1996;
Marschner and
Bredow, 2002
Andersson and Nilsson,
2001; Ogawa et al., 2001
Hongve, 1999
d
;
Hongve et al., 2000
d
;
Søndergaard et al., 2000
CO
2
,O
2
inoculum Jones and Edwards, 1998 Gilbert, 1988;
Amon and Benner,

1994
e
, 1996
e
Møller et al., 1999
f
Kalbitz et al., 2003
inoculum +
nutrients
Lundquist et al., 1999 ISO 10707 (1994);
Amon et al., 2001
d
Jandl and Sletten, 1999;
Jandl and Sollins, 1997;
Moran et al., 2000
Flow-through
reactor
DOC inoculum Volk et al., 1997;
Yano et al., 1998, 2000;
Søndergaard et al., 2000;
Søndergaard and Worm, 2001
In soil DOC – Brunner and
Blaser, 1989
Merckx et al., 2001;
Marschner and Noble,
2000; Marschner and
Bredow, 2002
Boudot et al., 1989
In soil CO
2

Jones and Edwards, 1998;
Jones et al., 2001;
Stro¨m et al., 2001
a
TOC: Total organic carbon (analysis without filtration).
b
Biologically active sand was used as an inoculum.
c
Biologically active sand was used as an inoculum besides suspended inocula.
d
No direct addition of inoculum; use of unfiltered samples (Zsolany and Steindl, 1991 used 1-Am filters).
e
Besides O
2
consumption, TOC concentration was measured.
f
Quantification of DOM biodegradability by measuring the denitrification potential.
B. Marschner, K. Kalbitz / Geoderma 113 (2003) 211–235216
this fract ion must consist of structures that are not easily cleave d by enzymes, such as
lignin degradation products or compounds strongly altered through preceding degradation
steps (Joergensen, 1998).
As pointed out by Qualls and Haines (1992) and Kalbitz et al. (2003), soil solutions
contain very different amoun ts of these fractions, and consequently, the kinetics of the
degradation process will be very different. Quantificat ion of the contribution of DOM
to the stable C pool in the mineral subsoil requires a quantification of rapidly and
slowly degradable DOM fractions and thei r mean residence times. Knowledge about
the size of the biodegradable DOM fraction is not sufficient. Kalbitz et al. (2003)
reported two DOM solutions with a similar portion of biodegradable DOC, whereas the
decomposition constants of the rapidly and slowly degradable fraction differed to a
great extent.

The other aspect that can strongly influence the result of a biodegradation assay is how
DOM degradation is quantified. If the change in DOC concentration is used, then samples
taken from the incubation solution will need to be filtered at 0.45 Am. Any DOC
transformed into microbial biomass or other particulate carbon resulting from coagulation
and precipitation will be largely retained on the filter and its disappearance will therefore
falsely be interpreted as degradation. On the other hand, if samples are left unfiltered and
TOC is used as a measure, microbially incorporated DOC will be regarded as not
degraded. According to Søndergaard et al. (2000), microbial biomass may be in the range
of 10% of TOC, so that the error made with TOC or DOC measurements may be small.
However, the error could be much higher if all particulate C is considered (C which does
not pass through a 0.45-Am filter). Thus, Kalbitz et al. (2003) reported a DOC mine-
ralisation after 90 days of only 9% from CO
2
data, although the DOC content of this
sample declined by 50%. However, a problem with TOC analysis is that an adequate and
reproducible sampling of the suspension and a complete combustion of the particles
cannot be guaranteed.
The other measure for DOM mineralisation is CO
2
efflux from the samples (Table 1).
However, even with this method, errors can occur due to CO
2
dissolution in water (3.4 g
CO
2
/l under atmospheric pressure in distilled water) if CO
2
is not trapped and depleted
from the atmosphere of the incubation vessel and when the solution was not in equilibrium
with atmospheric CO

2
initially.
Another approach for the determination of DOM degradability is using ‘‘bioreactors’’
filled with glass beads that are colonized by microorganisms to form so-called biofilms on
their surfa ces (Yano et al., 1998; Søndergaard et al., 2000). DOM solutions are passed
through such flow-through reactors and DOM degradation is determined from the
difference in DOC concentrations between in- and out-flow, usually with residence times
of 10–24 h. In a comparative study, Søndergaard et al. (2000) showed that the
degradability of DOM determined with such a system is closely related to DOM
degradability in batch cultures after 135–151 days (r
2
= 0.73) and reaches about 90% of
the batch values. This high efficiency of the bioreactor can be explained by the relatively
high microbial density compared to batch cultures which allows more intensive microbial
interactions with DOM and its degradation products within the biofilm. Pinney et al.
(2000) describe another type of bioreactor where they used biologically active sand in a
batch vessel.
B. Marschner, K. Kalbitz / Geoderma 113 (2003) 211–235 217
However, these flow-through bioreactors initially require long equilibration times (up
to 6 months) and continuous maintenance to achieve reproducible degradation rates
(McDowell, personal communication). Other problems of these reactors are the release of
DOM into the solution and the adsorption of DOM onto the biofilms. Furthermore, it
seems impossible to determine the pool sizes and the residence times of a rapidly and a
slowly degradable fraction.
The other listed parameters will also affect the result of DOM biodegradation measure-
ments because these are mainly discussed under soil and solution properties (Section 5.2).
Here are some examples. Shaking of DOM solutions during the incubation could hinder
the development of hyphae which could result in an underestimation of the biodegradation
by fungi. Addition of nutrients will at least accelerate DOM biodegradation resulting in
higher degradation rates in comparison to incubation without addition of nutrients

(Schmerwitz, 2001). Furthermore, an enhanced coagulation and precipitation due to
increased ionic strengths (Kalbitz et al., 2000) is imaginable. Finally, data about the
density of added mic roorganisms are scarce in published studies on DOM biodegradation,
ranging from 1000–2000 (Miettinen et al., 1999) to 10
4
CFU (Volk et al., 2000), and
0.48
Â
10
5
cells ml
À 1
(Buffam et al., 2001) in the inoculated sample. Mostly, it is reported
that 1% (v/v) of inoculum was added.
4. Controls of DOM bioavailability
The bioavailability of DOM is reduced if the possibilities of microorganisms to interact
with DOM are restricted. These may be physical restrictions, such as inaccesibility of
DOM in very small pores or chemical restrictions, such as DOM sorption to solid surfaces.
4.1. Pore size
DOM in small pores is not accessible for microorganisms, i.e., in pores with diameters
below 0.2 Am (Zsolnay, 1997). This pore size class contains water that is not plant
available and hardly participates in transport processes. Consequently, DOM in these pores
will only become bioavailable through diffusion into larger water-filled pores. Although
enzymes excret ed by microorganism may enter these pores, this is also limited to
diffusion, as well as the movement of the breakdown products out of the pores. In some
clayey soils, up to the 50% (v/v) of the total pore volume is in this size class and DOM
could be preserved there from microbial breakdown. However, to date, little is known
about the amounts and quality of DOM in different pore size classes, because the soil
water in the smallest pore sizes cannot be directly extracted for analysis.
Zsolnay and Steinweg (2000) have attempted to overcome this problem by using a

stepwise extraction technique to obtain DOM fractions from different pore size classes. In
a first step, the undis turbed samples are percolated to obtain the so-called mobile DOM.
Soil solution from the mesopores (0.2–6 Am) is then extracted by centrifugation, and the
remaining DOM is extracted in a batch-shake procedure, with a mild salt solution. For the
three soils examined by Steinweg (2002), DOM in the percolates was always the least
biodegradable, thus supporting the assumption that this DOM pool should be depleted first
B. Marschner, K. Kalbitz / Geoderma 113 (2003) 211–235218
due to its high bioavailability. In the other two fractions, DOM degradability was similar
and much higher, thus indicating some physical protection in less-accessible pores. Since
71–82% of DOM was obtained in the batch extracts, DOM conservation in the smallest
pores may be of major importance in soils.
4.2. Soil aggregation
A similar mechanism of restricted bioavailab ility of DOM in small pores may occur
within aggregates since various studies have shown that the disruption of aggregates
stimulates microbial activity (i.e., Elliott, 1986; Ladd et al., 1993) and that aggregates
contain more young and less-altered plant-derived organic matter than the bulk soil
(Skjemstad et al., 1990; Puget et al., 1995; Six et al., 2000). However, no studies have
been encountered where DOM from within aggregates was compared to bulk soil DOM in
terms of its biodegradability.
4.3. Sorption
In the presence of mineral solid phases, the mineralisation of plant-derived carbohy-
drates or simple organic compounds such as glucos e and citrate can be greatly reduced,
especially if charged molecules like citrate or oxalate interact with charged minerals such
as clay minerals or goethite (Jones and Edwards, 1998; Miltner and Zech, 1998; Stro¨m
et al., 2001). The observed reduced biodegradability of soil organic matter through
sorption to mineral surfaces is considered to be one or the most important stabilisation
processes, and it is extensively reviewed by Sollins et al. (1996) and Kaiser and
Guggenberger (2000). However, the mechanisms of this sorption process are still as
poorly understood, as the reasons why sorbed materials may be less degradable. In
contrast, Guggenberger and Kaiser (2003) estimated a mean residence time of the sorbed

organic carbon of about 4–30 years and thus challenged the commonly assumed
sorptive stabilization of DOM. They hypothesised that natural soil surfaces are covered
by biofilms with a high affinity for DOM, so that the observ ed ‘‘sorption’’ may indeed
enhance bioavailability and subsequent biodegradation. Only sorption onto purely
mineral surfa ces would thus result in an effective stabilization of DOM (Guggenberger
and Kaiser, 2003). On the other hand, sorption is generally not irreversible, so that
sorbed materials may return to the solution phase and thus become bioavailable again.
Desorption will depend on the nature of the sorbate and sorbent and is affected by
solution composition. Kaiser and Guggenberger (2000) have shown that the hydrophilic
DOM fraction is much more easily desorbed than the hydrophobic DOM fraction, which
is also less biodegradable (see below).
4.4. Drought
DOM availability for microorganisms is reduced when soils become dry, since this
greatly limits microbial activity and decreas es diffusive transport processes towards the
remaining moist sites of activity. The strong stimulation of microbial activity after
rewetting dry soils is therefore often attributed to the accumulation of easily degradable
B. Marschner, K. Kalbitz / Geoderma 113 (2003) 211–235 219
substances such as cellular materials from the desiccated organisms in the dry soil
(Lundquist et al., 1999; Zsolnay et al., 1999; Merckx et al., 2001).
5. Factors controlling DOM biodegradability
The biodegradability of DOM is controlled by numerous factors that can be divided
into three categories. The first set of factors are intrinsic DOM characteristics that are
determined by molecular structure, functional group content or size of the molecules. The
second set of factors consists of soil properties that can influence the degradation process,
such as nutrient availability, microbial community structure and the presence of toxic
substances or other soil solution constituents. At the third level, external factors such as the
temperature and rainfall regime and the associated vegetational cycles will induce a
seasonal variability of both DOM inputs and microbial activity which can affect intrinsic
DOM quality parameters and soil and soil solution properties.
5.1. Intrinsic DOM quality parameters

5.1.1. Molecular size
Considering the uptake mechanisms of microorganisms, one could expect that smaller
DOM molecules or units should be ingested and degraded preferentially. Evidence for this
was found in one of our studies where the biodegradability of DOM in ultrafiltrates of the
size class < 1000 Da was three- to fourfold higher than in the size class < 10,000 Da or in the
bulk DOM solution (Table 2). However, this was only true for DOM extracted from soil
samples that were collected in early spring. In summer, biodegradability of DOM was much
lower, with no differentiation between size classes. The reason for this is probably the
depletion of degradable compounds by the activated microorganisms during late spring and
summer. On the other hand, the preferential degradation of small compounds in the spring
sample may not be a size effect but due to chemical characteristics. Kaiser et al. (2001b)
showed for a forest soil that easily degradable carbohydrates, amino sugars and proteins
accumulate during winter and these compounds would largely appear in the small size class.
For aquatic DOM, Amon and Benner (1994, 1996) found opposite results. In their
samples obtained from the Gulf of Mexico and the Amazon River and nearby coastal
ocean waters, they determined a much higher C mineralisation from larger DOM size
faction (>1000 Da) compared to smaller DOM. Since most DOM in the Amazon was in
the larger size fraction and marine DOM consisted mainly of the small size fraction, they
Table 2
Effect of sampling date on the biodegradability of total DOM and DOM in two size fractions (ultrafiltration) from
solutions obtained from percolating undisturbed soil samples from an arable field with 0.01 M CaCl
2
(DOC after
5-day incubation at 20 jC in % of initial DOC)
Sampling date Total DOM DOM < 10 kDa DOM < 1 kDa
March 12.1 a 16.2 a 51.3 b
July
3.3 a 3.5 a 4.4 a
Values in rows followed by the same letter are not significantly different ( p < 0.05, Duncan test).
B. Marschner, K. Kalbitz / Geoderma 113 (2003) 211–235220

concluded that the larger molecules are more bioreactive not due to their size, but because
they are fresher, i.e., less degraded than the smaller compo unds. Therefore, again size is
only a secondary attribute and the primary factor controlling DOM biodegradability would
be structural characteristics.
5.1.2. Chemical structure and spectroscopic properties
Carbohydrates and amino acids are highly decomposable in soils (Haider, 1992) and are
utilized preferentially by microorganisms during degradation of different compounds in
DOM solutions (Volk et al., 1997; Amon et al., 2001; Kalbitz et al., 2003). However, Volk
et al. (1997) stated that the often used classification of carbohydrates as labile DOM
components should be seen with caution, as carbohydrates can also be bound to stable
DOM compounds.
Compounds with alkyl or aromatic structural units generally accumulate during the
decomposition of soil organic matter (Baldock et al., 1992; Ko¨gel-Knabner et al., 1992;
Baldock and Preston, 1995; Huang et al., 1999) and have thus been associated with a low
biodegradability. Boissier and Fontvieille (1993) found that phenols and polyphenols were
closely related to the amount of nondegradable DOM in incubation experiments. Similarly,
Wershaw and Kennedy (1998) observed a relative increase in aromatic structures during
litter decomposition. Kalbitz et al. (2003) showed that the biodegradability of DOM
extracted from forest litter layers was negatively correlated to its content in aromatic
structures determined with
1
H-NMR.
Spectroscopic properties are commonly determined for DOM characterisation. Traina et
al. (1990) and Chin et al. (1994) have shown that the specific UV absorbance of humic and
fulvic acids between 250 and 280 nm is closely correlated to their content in aromatic
structures.
Since aromatic structures are generally quite recalcitrant, one would expect a negative
relationship between the specific UV absorbance and the biodegradability of DOM. This
indeed was observed in one of our studies (Fig. 2a,Jo¨demann unpublished results) with a
linear correlation coefficient of r = 0.69 for 28 solutions obtained from an arable soil that

had been stor ed fresh, air-dried or frozen and then extracted with 1 mM CaCl
2
solution
with a percolation procedure with either undisturbed samples or after homogenization. In
the same samples, the decrease in DOC concentration after biodegradation was highly
significantly correlated (r = 0.85) with an increase in specific UV absorbance (Fig. 2b),
thus indicating that UV-inactive substances were degraded preferentially. Other authors
also reported close correlations between DOM degradability and specific UV absorbance
(Gilbert, 1988; Zoungrana et al., 1998; Pinney et al., 2000; Kalbitz et al., 2003) and some
even found nonlinear relationships, where biodegradability incre ases exponentially with
decreasing UV absorbance.
However, specific UV absorbance of DOM is not always a reliable predictor for
biodegradability. Marschner and Bredow (2002) show that the biodegradabi lity of DOM
from soil samples incubated at different temperatures varied greatly from 8% to 61% but was
not related to specific UV absorbance of either total DOM or of its size fractions. If one
accepts the assumption that UV absorbance is a measure for the recalcitrant aromatic
structures, then these results clearly show that the non-aromatic compounds also greatly
differ in biodegradability. A low biodegradability of aliphatic compounds may be due to
B. Marschner, K. Kalbitz / Geoderma 113 (2003) 211–235 221
binding to aromatic structures (i.e., lignocellulose) or to a high degree of polymerisation or
oxidation (Guggenberger et al., 1994), but this cannot be assessed with simple spectroscopic
methods.
More recently, fluorescence spectroscopy has been used successfully to obtain
information about the biodegradability of DOM (Glatzel et al., 2003; Kalbitz et al.,
2003) using the assumption that more condensed aromatic stru ctures with a red-shifted
fluorescence are less biodegradable than structures with a low degree of condensation and
conjugation. Zsolnay et al. (1999) showed that a humification index calculated from
fluorescence data can help to differentiate between microbial cell lysis products and more
humified DOM. Parlan ti et al. (2000) stressed the usefulness of fluorescence spectroscopy
as an indicator for biological activity and humification in coastal waters.

Fig. 2. Relationship between DOC degradation and specific UV absorbance of DOC extracted with 0.01 M CaCl
2
solution from differently treated top soil samples from an arable field. Sample treatments included drying and
freezing prior to extraction. (a) Relationship between initial UV absorbance and degradation of DOC after 5 days
of incubation. (b) Relationship between the change in UV absorbance during incubation and the degradation of
DOC.
B. Marschner, K. Kalbitz / Geoderma 113 (2003) 211–235222
5.1.3. Fractionation according to polarity and acidity
A DOM fractionation scheme used frequently is the one developed by Leenheer (1981),
where DOM is separated according to its polarity and acidity. Jandl and Sollins (1997)
performed incuba tion studies with such different DOM fractions obtained from water
extracts of forest soil samples. For the acidic fractions (hydrophilic and hydrophobic),
degradation was extremely low ( < 5%) during 120 days of incubation. In contrast, the
hydrophilic neutral fraction was min eralised to almost 15%, mostly during the first 3 days.
Since this fraction is enriched with carbohydrates from cellulose and hemicellulose
breakdown and from microbial origin (Guggenberger et al., 1994), its high biodegrad-
ability is probably due to these easily utilisable substances. As a consequence, this fraction
generally amounts to less than 20% of total DOC in soil solutions (Vance and David, 1991;
Qualls and Haines, 1992; Guggenberger and Zech, 1993; Guggenberger et al., 1994;
Andersson and Nilsson, 2001; Kaiser et al., 2001b) . In consequence, the hydrophilic and
hydrophobic acid fractions generally dominate, which is attributed to their content of
recalcitrant compounds and their higher degree of biodegradation (Guggenberger et al.,
1994). Kalbitz et al. (2003) found a close negative correlation between DOM degradability
and hydrophobic DOM portions.
The hydrophobic acid fraction is preferentially sorbed to mineral surfaces, especially Fe
and Al oxyhydroxides (Dai et al., 1996; Kaiser and Guggenberger, 2000).Asa
consequence, the hydrophilic fraction becomes more dominant with increasing depth
and decreasing DOC concentrations, so that biodegradability of DOM may increase, as
observed by Qualls and Haines (1992).
5.2. Soil and solution properties

Soil properties that affect the physical and chemical environment of the microbial
degrader community are expected to affect their activity, and therefore, the degradation of
DOM in situ. Furthermore, intrinsic DOM properties are affected by soil and solution
properties. However, most studies of DOM biodegradability are conducted in vitro, i.e., in
solution cultures without the soil solid phase. Therefore, only effects of varying solution
parameters on DOM biodegradability can be assessed.
5.2.1. Nutrients, salts, pH, O
2
A question that is addressed repeat edly in these studies concerns the addition of
inorganic nutrients to the bioassays. If the potential biodegradability of DOM is to be
assessed, all other limits to microbial activity should be eliminated, i.e., macro- and
micronutrient suppl ies optimized. This is generally restricted to addition of N, P and K,
since it is assumed that other nutrients are generally present in adequate amounts in the soil
solutions or soil extracts. In a study with five selected DOM solutions extracted from
forest floors, peat and agricultural soils, Schmerwitz (2001) found either no effects of NPK
additions on DOM biodegradability or only minor stimulating or depressing effects (Fig.
3). Enhancement occurred with DOM of low degradability, while slightly negative nutrient
effects were observed for the highly degradable DOM samples. Nelson et al. (1994)
determined the e ffects of N additions on the mineralisation of DOM obtained from water
extracts of soil samples taken at different depths of a pasture profile. In the subsoil (80–
B. Marschner, K. Kalbitz / Geoderma 113 (2003) 211–235 223
100 cm), DOM degradability was not N limited, but in the topsoil, N additions increased
CO
2
release from the DOM solutions by roughly 25%. Nelson et al. (1994) attribute this to
a relatively high supply of fresh root exudates and soluble root compounds that are easily
degradable, so that biodegradability becomes N limited. This appears to be an exception,
since microbial growth in soils is generally limited by the availability of C substrates
(Lynch, 1982), and therefore mainly stimulated by inputs of fresh C sources.
However, at this point, it may be helpful to consider an important aspect of substrate

utilisation by m icroorganisms. All heterotrophic organisms need C sources for two
purposes. One is as an energy substrate where organic carbon is largely mineralised to
CO
2
. This form of C utilisation mainly depends on the size of the microbial population,
and therefore microbial biomass can be estimated from the CO
2
release after glucose
addition with the substrate-induced respiration (SIR) method of Anderson and Domsch
(1978). With these pure energy substrates, the second purpose, microbial growth, will not
occur unless important nutrients like N and P are present that are needed for the synthesis
of new biomass. Therefore, nutrient limitations for the degradation of DOM will only arise
when the ratio of biodegradable DOM to microbial biomass is too wide, and when DOM is
poor in N or P. In soils, this is probably rarely the case. However, in solution bioassays that
need inoculation, microbial growth may be a prerequisite for DOM degradation and
therefore may be stimulated by nutrient additions.
In their incubation experiments with different DOM fractions, Jandl and Sletten (1999)
also tested the effects of Ca on DOM degradation. At 1:1 molar ratios of Ca/DOC, the
degradation of hydroph obic acids from litter extracts and soil solutions was increased 1.5-
to 6-fold compared to a Ca-free control. For the hydrophilic fractions, Ca had a slightly
inhibitory effect and even glucose was degraded less in solutions containing Ca. Since
glucose cannot form stable complexes with Ca, the authors suggest that Ca forms
complexes with metabolites and thu s stabilis es them against further biodegradation.
Similar mechanisms may be active with the hydrophilic fractions. The stimulatory effects
Fig. 3. Effects of nutrient addition (N + P + K) on the degradation of DOC from different forest floor solutions
during a 10-day incubation (adapted from Schmerwitz, 2001).
B. Marschner, K. Kalbitz / Geoderma 113 (2003) 211–235224
of Ca on the degradation of the hydrophobic acid fraction are assumed to be due to
flocculation, as larger structures will provide attachment for microbial colonies (Jandl and
Sletten, 1999). However, such precipitation of DOM could also result in a decreased

bioavailability because DOM is removed from the aqueous phase. In soils, Ca additions
generally reduce DOM solubility, affecting mainly the large size fraction (Ro¨mkens and
Dolfing, 1998). It has not yet been established if this has positive or negative effects on the
degradation of the sorbed or the remaining soluble DOM.
Other solution parameters such as O
2
, pH and soluble salts will also influence microbial
activity and the solubility and configuration of DOM molecules. Under anoxic conditions,
DOC concentrations are commonly elevated (Hunchak-Kariouk et al., 1997; Hage dorn et
al., 2000) , which is either attributed to the absence of oxide sorption sites or to the great ly
reduced microbial activity so that easily degradable organic compounds like acetate may
accumulate in the soil solution (Ku
¨
sel and Drake, 1999). At low pH, DOM solubility is
generally lower and molecules are more condensed than at higher pH, while Na
+
or K
+
can
increase DOM solubility and cause an expansion of DOM molecules in contrast to Ca
effects (Ghosh and Schnitzer, 1980; Murphy et al., 1994). Since most of these solution
parameters will also directly affect the composition a nd activity of the microbial
community (Metting, 1993), it should be difficult to separate these effects from those of
configurational changes on biodegradability.
5.2.2. Metal concentration
While nutrient availability in the soil solution can affect DOM degradability through its
effects on microbial activity, other solution components can directly interact with DOM
and thus may alter its biodegradability. In acid forest soils, Al and Fe can form relatively
stable complexes with DOM and may thus be mobilized and transported in the soil profile,
as it is observed during podsolisation (Blaser, 1994). When these complexes form, DOM

is altered stru cturally, as evidenced by changes in fluorescence (Blaser et al., 1999) and
molecular size (R itchie and Posner, 1982; Jandl and Sletten, 1999). Several authors have
assumed that DOM in these complexes is stabilised against biodegradation due to toxic
effects from the metals, especially Al (Brunner and Blaser, 1989; Jones et al., 2001).
However, experimental data for this are scarce and conflicting. Jandl and Sletten (1999)
observed inhibiting, enhancing and no effects of Al additions on the mineralisation of
different DOM fractions from forest floor solutions. Since the strongest inhibiting effects
were observed for the carbohydrate-rich hydrophilic neutral fraction and for glucose, direct
Al toxicity is more likely since these uncharged molecules are poor metal complexers.
Jones et al. (2001) found no Al effects on the mineralisation of organic acids unless Al
concentrations were increased to 5 mM, where they were equally effective in reducing the
mineralisation of 0.5 mM citrate or oxalate. Boudot et al. (1989) found inhibiting effects of
Al and Fe on citrate mineralisation only at molar ratios of at least 0.97 and suspect that this
is likely due to precipitation of metal oxides and subsequent sorption of their organic
model substance. On the other hand, Lundstro¨m et al. (1995) showed that the mineralisa-
tion of organic metal complexes may be a prerequisite for podsolisation. They percolated
litter extracts and organic acids through columns filled with C material and observed the
formation of an eluvial horizon depleted in Al and Fe after 10 months. However, an Al-
and Fe-enriched illuvial horizon only was formed in the biologically active columns where
B. Marschner, K. Kalbitz / Geoderma 113 (2003) 211–235 225
the mineralisation of the complexes caused the precipitation of the metals. In the sterile
treatments, no Al or Fe deposition was observed within the columns.
These reports indicate that the complexation of potentially toxic metals by DOM will
not necessarily reduce its biodegradability but instead may even enhance microbial activity
by reducing the free metal ion concentrations and thus their toxicity. This would be similar
to the beneficial effects of DOM on Al toxicity to roots and microalgae (Marschner, 1995;
Parent et al., 1996).
Certain heavy metals such as Cu, Pb and Hg can also form very stable complexes with
DOM which can alter metal toxicity to aquatic organisms (Oikari et al., 1992; Alberts et
al., 2001) and the structural composition of DOM (Blaser et al., 1999). However, no

studies were encountered where the degradability of metal complexes was compared to
that of metal-free DOM. Inhibitory effects of heavy metals on microbial activity in soils
have been studied extensively and this has been reviewed recently by Giller et al. (1998).
In a very original experimental setup, Merckx et al. (2001) showed that Zn additions to
soil ranging from 50 or 500 mg/kg inhibited the degradation of DOM that had been
released after rewetting the dried soil. However, the authors interpret this as a direct toxic
effect rather than a stabilization of DOM by the complexed metals.
5.2.3. Organic compounds
As mentioned above, aromatic DOM compounds are generally more stable than
molecules with aliphatic structures. In addition to this, soluble polyphenols, phenolic
acids and plant-derived tannins have been shown to inhibit the activity of various enzymes
(Benoit et al., 1968; Williams and Gray, 1974; Gianfreda et al., 1995; Wetzel, 2000). The
inhibitory effects of tannins on enzyme activity are less pronounced in the presence of
polyvalent cations such as Al
3+
,Fe
3+
or Mn
2+
(Gianfreda et al., 1995), which may also
explain the stimulatory effects of Al additions on the degradation of the forest soil-derived
hydrophobic acid fraction mentioned above (Jandl and Sletten, 1999).
Other natural organic compounds that may even be toxic to soil microorganisms and
can thus affect degradation processes include terpenoids (Bremner and McCarty, 1993)
and certain amino acids like mimosine (Soedarjo and Borthakur, 1998). Fritze et al. (1998)
report that DOM extracted from burned forest floor greatly reduced CO
2
release when
added to soil samples. They found the highest concentration of toxic DOM in the
hydrophilic base fraction but were not able to identify the responsible compounds.

5.2.4. Composition of the microbial community
From a general ecological and evolutionary understanding, one would expect the
autochthonous microbial population to be best adapted to the optimal utilisation of the
organic compounds present in a certain soil. This hypothesis was proven by Block et al.
(1992), who tested the effect of different inocula on the biodegradation of aquatic DOM.
They concluded that indigenous mixed bacterial populations shoul d be used to determine
the biodegradability of DOM. However, land use changes and introduction of different
plant species alters the quality and quantity of organic matter entering the soil system
(Sanger et al., 1997; Bauhus et al., 1998; Coudron and Newman, 1998; Chen and Stark,
2000), which should also alter the quality of DOM. In order to assess the influence of
different microbial communities on DOM degradation, Schmerwitz (2001) inoculated four
B. Marschner, K. Kalbitz / Geoderma 113 (2003) 211–235226
DOM solutions with soil extracts obtained from these four different sites (beech fores t,
spruce forest, peat and agricultural soil) and with a mixed inoculum containing the
microorganisms from all sites. The degradability of the DOM solutions differed by up to a
factor of two depending on the origin of the inoculum. However, the highest DOM
degradation was not observed with the native inocul um, and noninoculated samples also
showed high DOM degradations. Although t he reasons for the different inoculum
efficiencies are not known, these results clearly show that the assessment of DOM
biodegradability depends on the type and origin of the microorganisms used for the test.
Møller et al. (1999) shed some light on the complexity of microbial interactions that
may occur during these degradation processes. They incubated sterilized beech litter with
different cultures of bacteria, fungi and c ombinations of both and determined the
biodegradability of extractable DOM by determining its denitrification potential (DNP).
As expected, the DNP of DOM from the sterile leaves was highest, because the fresh plant
components had not been utilised by microorganisms yet. After incubation of the leaves
with bacteria alone or with bacteria + fungi, DNP of extractable DOM was lower, but the
lowest DNP was determined for DOM extracted after incuba tion with a cellulytic fungus
(Humicola sp.), especially in the presence of bacteria which apparently very effectively
removed a large p ropo rtion of the biodegradable DOM from the solutio n through

mineralisation and incorporation into the microbial biomass. This is a nice example for
the important interactions of different organisms during the degradation of DOM.
5.3. External factors
The DOM characteristics and soil properties controlling DOM biodegradability are not
only site-specific characteristics, but they will also vary in time due to seasonal changes in
OM inputs, temperature and moisture regime. DOC concentrations in soil solutions are
generally highest during summer (Hongve, 1999; Kalbitz and Popp, 1999; Kaiser et al.,
2001b; Yano et al., 2000) whi ch is attributed to the combined effects of increased release
of root exudates and microbial metabolites. Concentration effects due to reduced water
content can also occur, but are of minor magnitude. Wet – dry cycles during summer can
also contribute to elevated DOM concentrations, due to aggregate disruption, microbial
cell lysis and stimulated microbial activity (Zsolnay and Go¨rlitz, 1994; Borken et al.,
1999; Lundquist et al., 1999).
Only few data are available on the seasonal variability of DOM quality or biodegrad-
ability. Kaiser et al. (2001b) showed from
13
C-NMR analyses that DOM in forest floor
leachates contained more low-molecular organic acids and less aromatic, O-alkyl struc-
tures and COOH groups during winter than in summer. From this, they conclude that
winter DOM should be more degradable than summer DOM and that this is due to the
accumulation of easily degradable compounds during the dormant season . These findings
agree well with other studies which generally observe higher DOM degradabilities in
winter/spring than in summer/fall (Qualls and Haines, 1992; Nelson et al., 1994; Hongve,
1999; Lundquist et al., 1999). Howe ver, Yano et al. (2000) determined DOM biodegrad-
abilities as high as 40% in summer soil solutions while during winter, degradability was
only 10–20%. Since the nondegradable DOM concentrations remained fairly stable during
the year, they attributed the observed seasonality of easily degradable DOM primarily to
B. Marschner, K. Kalbitz / Geoderma 113 (2003) 211–235 227
the release of organic compounds from roots. Another source of easily degradable DOM
during the growing season could be canopy-derived plant leachates.

Within the physiological range of 0 – 35 jC, temperature is known to stimulate
microbial activity (Paul and Clark, 1996). As a consequence, DOM production may
increase at higher temperatures due to enhanced microbia l breakdown of larger insoluble
compounds to soluble entities, as observed by Christ and David (1996). On the other hand,
increased microbial activity also results in enhanced biodegradation and mineralisation of
DOM. If this process dominates over the enhanced production, the net effect of a
temperature-stimulated microbial activity will be DOM depletion. This was observed by
Marschner and Br edow (2002), where extra ctab le DOM decr ease d with incr ea sing
temperature, while at the same time DOM became more biodegradable (Table 3). This
stands in contrast to the concept that easily degradable DOM should be depleted
preferentially in soils. One possible explanation is that the high microbial activity caused
a rapid depletion of substrates and nutrients , resulting in the die-back of part of the
microbial population, releasing easily degradable cell constituents into the soil solution. In
any case, these studies show that the external controls on DOM production and DOM
quality, especially in terms of biodegradability, are only poorly understood.
6. Conclusions
This review shows that the biodegradability of DOM in soil solutions or in aqueous soil
extracts varies greatly with soil depth, samp ling date or site characteristics. Some of the
differences can be attributed to the presence of varying amounts of specific chemical
compounds like sugars, proteins, phenols or tannins that are known for their different
biodegradability. However, as with other soil organic matter, DOM contains only a small
fraction of these identifiable compounds, while the major part are structures that have been
altered by microbial and biochemical degradation processes. The degrada bility of these
compounds will be mai nly determined by the presence of structural components for which
enzymatic tools exist within the microbial c ommunity. Little is known about the diversity
and efficiency of these enzymes, so that it seems unlikely that an exact characterisation of
structural components will allow the prediction of DOM degradability. However, the
correlations between biodegradability and certain DOM characteristics, such as specific
UV absorbance, aromaticity, hydrophilic neutral or hydrophobic acid content show that
certain analytical tools can help to explain differences in DOM degradability. However,

Table 3
Effects of soil sample incubation temperature on DOM extractability (percolation with 0.01 M CaCl
2
) and on
DOM degradability (DOC after 5-day incubation at 20 jC in % of initial DOC)
Before incubation After incubation at
5 jC20jC35jC
DOC [mmol kg
À 1
] 26 a 18 b 7 c 3 d
DOC degradation [%] 12 a 8 a 48 b 61 b
Values in rows followed by the same letter are not significantly different ( p < 0.05, Duncan test). Data from
Marschner and Bredow (2002).
B. Marschner, K. Kalbitz / Geoderma 113 (2003) 211–235228
conclusions about causal relationships cannot be easily drawn from these results. For
example, the generally negative correlation between UV absorbance as a measure for
aromaticity and degrada bility can be due to the recalcitrance of aromatic structures or due
to inhibitory effects of these compounds on enzyme activity.
Among the studies investigating other factors that may influence DOM degradability,
results are often conflicting or ambiguous so that not even concepts about causalities can
be developed yet. This concerns the effects of nutrients or metals on biodegradability,
while other potential controlling factors such as pH or salts have not been assessed at all. A
summary of the discussed controlling factors is presented in Fig. 4. Here also, some
parameters have been included, which may be of relevance although no experimental
evidence for this has been encountered.
The major obstacle for a better understanding of the contr ols of DOM biodegradability
is the lack of a standardised methodology or at least systematic comparisons between the
various methods used to assess DOM biodegradability. The high variability in incubation
durations, inoculum or nutrient additions and the different measures for the quantification
of DOM degradation greatly hinder comparisons between the studies. Therefore, efforts

should be put in the development of a standardised protocol for DOM degradation studies.
For this, interlaboratory comparisons have to be made that assess the variability of
degradabilities determined with the different methods on the same samples. A first such
study is currently under way among six laboratories in Europe and North America.
Methodological research should also include the specific features of soils. Almost
nothing is known about the effects of natural soil surfaces with a high density and diversity
of microorganisms on DOM biodegradation.
Finally, the question of bioavailability has been addressed in only very few studies. In
naturally structured soils, certain amounts of DOM may not be accessible for micro-
Fig. 4. Summary of the three groups of parameters that have been identified as controlling factors for DOM
degradability. Bold: verified in several studies; italic: with conflicting or circumstantial evidence in some studies
or assumed factor.
B. Marschner, K. Kalbitz / Geoderma 113 (2003) 211–235 229
organisms. Soil solutions collected in situ originate only from larger pores, while batch
extractions disrupt soil structures to such a degree that solubilisation of previously solid
organic matter may occur. Therefore, it is still an unsolved methodological challenge to
obtain, quantify and characterise this fraction.
Acknowledgements
We would like to thank two anonymous reviewers and Adam Zsolnay for their very
helpful comments and criticism.
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