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

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-
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.

– 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

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,

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.

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

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

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

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.

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

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

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.

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.

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

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