Advances in Agronomy Volume 110

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V O L U M E O N E H U N D R E D T E N
ADVANCES IN
AGRONOMY
ADVANCES IN AGRONOMY
Advisory Board
PAUL M. BERTSCH
University of Kentucky
RONALD L. PHILLIPS
University of Minnesota
KATE M. SCOW
University of California,
Davis
LARRY P. WILDING
Texas A&M University
Emeritus Advisory Board Members
JOHN S. BOYER
University of Delaware
KENNETH J. FREY
Iowa State University
EUGENE J. KAMPRATH
North Carolina State
University
MARTIN ALEXANDER
Cornell University
Prepared in cooperation with the
American Society of Agronomy, Crop Science Society of America, and Soil
Science Society of America Book and Multimedia Publishing Committee
DAVID D. BALTENSPERGER, CHAIR
LISA K. AL-AMOODI CRAIG A. ROBERTS
WARREN A. DICK MARY C. SAVIN
HARI B. KRISHNAN APRIL L. ULERY
SALLY D. LOGSDON
V O L U M E O N E H U N D R E D T E N
ADVANCES IN
AGRONOMY
EDITED BY
DONALD L. SPARKS
Department of Plant and Soil Sciences
University of Delaware
Newark, Delaware, USA
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11 12 13 10 9 8 7 6 5 4 3 2 1
CONTENTS
Contributors ix
Preface xi
1. Dissolved Organic Matter: Biogeochemistry, Dynamics,
and Environmental Significance in Soils 1
Nanthi S. Bolan, Domy C. Adriano, Anitha Kunhikrishnan,
Trevor James, Richard McDowell, and Nicola Senesi
1. Introduction 3
2. Sources, Pools, and Fluxes of Dissolved Organic Matter in Soils 5
3. Properties and Chemical Composition of Dissolved Organic
Matter in Soils 13
4. Mechanisms Regulating Dynamics of Dissolved Organic
Matter in Soils 20
5. Factors Influencing Dynamics of Dissolved Organic Matter in Soils 30
6. Environmental Significance of Dissolved Organic Matter in Soils 37
7. Summary and Research Needs 60
Acknowledgments 62
References 62
2. Genomic Selection in Plant Breeding: Knowledge and Prospects 77
Aaron J. Lorenz, Shiaoman Chao, Franco G. Asoro, Elliot L. Heffner,
Takeshi Hayashi, Hiroyoshi Iwata, Kevin P. Smith, Mark E. Sorrells,
and Jean-Luc Jannink
1. Introduction 78
2. Important Population and Trait Characteristics 80
3. Single Nucleotide Polymorphism Marker Discovery
and Genotyping 82
4. Statistical Methods 84
5. GS Prediction Accuracies 94
6. Impact of Statistical Model on GEBV Accuracy 103
7. Modeling Epistasis and Dominance 107
8. GS in the Presence of Strong Subpopulation Structure 109
9. Long-Term Selection 111
10. Summary and Conclusions 114
References 116
v
3. Differences of Some Leguminous and Nonleguminous Crops in
Utilization of Soil Phosphorus and Responses to Phosphate
Fertilizers 125
Sheng-Xiu Li, Zhao-Hui Wang, and Bobby Alton Stewart
1. Introduction 130
2. The Difference of P Uptake Amounts of Leguminous and
Nonleguminous Crops 141
3. Leguminous and Nonleguminous Crop Responses to Powdered
Rock Phosphates 146
4. The Relation of Plants’ Root Morphology to Their Capacity of
Using Soil Sparingly Soluble P and Responses to P Fertilizers 163
5. Microorganisms in Rhizosphere Soil and Their Function
in Supplying P to Plants 173
6. Root Exudates (Substances Secreted from Roots) and the Plants’
Capacity to Use Sparingly Soluble P in the Soil and Crop
Responses to P Fertilizers 176
7. Effects of Root Cation Exchange Capacity and Calcium Uptake
Amount of Crops on Soil P Absorption and Crop Responses
to P Fertilizer 189
8. The Responses to P Fertilizer Between Leguminous and
Cereal Crops with Their Biological Characteristics 193
9. Factors Affecting the Responses of Leguminous and
Nonleguminous Crops to P Fertilizer 203
10. Conclusions 222
Acknowledgments 227
References 227
4. The Role of Mineral Nutrition on Root Growth of Crop Plants 251
N. K. Fageria and A. Moreira
1. Introduction 252
2. Root-Induced Changes in the Rhizosphere 255
3. Root Systems of Cereals and Legumes 256
4. Contribution of Root Systems to Total Plant Weight 260
5. Rooting Depth and Root Distribution 263
6. Root Growth as a Function of Plant Age 265
7. Root–Shoot Ratio 268
8. Root Growth Versus Crop Yield 270
9. Genotypic Variation in Root Growth 271
10. Root Oxidation Activity in Oxygen-Deficient Soils 274
11. Root Growth in Conservation Tillage Systems 276
12. Mineral Nutrition Versus Root Growth 278
13. Management Strategies for Maximizing Root Systems 312
vi Contents
14. Conclusions 317
Acknowledgment 318
References 318
5. Physiology of Spikelet Development on the Rice Panicle: Is
Manipulation of Apical Dominance Crucial for Grain
Yield Improvement? 333
Pravat K. Mohapatra, Rashmi Panigrahi, and Neil C. Turner
1. Introduction 334
2. Panicle Structure and Development 335
3. Panicle Architecture and Grain Yield 337
4. Physiological Factors Regulating Spikelet Development 340
5. Is Manipulation of Apical Dominance Crucial for Grain
Yield Improvement? 348
6. Suggestions for Modification of Apical Dominance 351
Acknowledgments 352
References 352
Index 361
Contents vii
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CONTRIBUTORS
Domy C. Adriano (1)
University of Georgia, Savannah River Ecology Laboratory, Drawer E, Aiken,
South Carolina, USA
Franco G. Asoro (77)
Department of Agronomy, Iowa State University, Ames, Iowa, USA
Nanthi S. Bolan (1)
Centre for Environmental Risk Assessment and Remediation (CERAR), and
Cooperative Research Centre for Contaminants Assessment and Remediation of
the Environment (CRC CARE), University of South Australia, Australia
Shiaoman Chao (77)
Biosciences Research Laboratory, USDA-ARS, Fargo, North Dakota, USA
N. K. Fageria (251)
Rice and Bean Research Center of Embrapa, Santo Antônio de Goiás, GO, Brazil
Takeshi Hayashi (77)
Data Mining and Grid Research Team, National Agricultural Research Center,
Tsukuba, Ibaraki, Japan
Elliot L. Heffner (77)
Department of Plant Breeding and Genetics, Cornell University, Ithaca,
New York, USA
Hiroyoshi Iwata (77)
Department of Agricultural and Environmental Biology, Graduate School of
Agriculture & Life Sciences, University of Tokyo, Bunkyo, Tokyo, Japan
Trevor James (1)
AgResearch, Ruakura Research Centre, Hamilton, New Zealand
Jean-Luc Jannink (77)
R.W. Holley Center for Agriculture and Health, USDA-ARS, Ithaca, New York,
USA
Anitha Kunhikrishnan (1)
Centre for Environmental Risk Assessment and Remediation (CERAR), and
Cooperative Research Centre for Contaminants Assessment and Remediation of
the Environment (CRC CARE), University of South Australia, Australia
ixx Contributors
Sheng-Xiu Li (125)
College of Resources and Environmental Sciences, Northwest Science and
Technology University of Agriculture and Forestry, Yangling, Shaanxi, PR China
Aaron J. Lorenz (77)
R.W. Holley Center for Agriculture and Health, USDA-ARS, Ithaca, New York,
USA
Richard McDowell (1)
AgResearch, Invermay Agricultural Centre, Mosgiel, New Zealand
Pravat K. Mohapatra (333)
School of Life Science, Sambalpur University, Sambalpur, India
A. Moreira (251)
Western Amazon Research Center of Embrapa, Manaus, AM, Brazil
Rashmi Panigrahi (333)
School of Life Science, Sambalpur University, Sambalpur, India
Nicola Senesi (1)
Department of Agroforestal and Environmental Biology and Chemistry, University
of Bari, Bari, Italy
Kevin P. Smith (77)
Department of Agronomy and Plant Genetics, University of Minnesota, St. Paul,
Minnesota, USA
Mark E. Sorrells (77)
Department of Plant Breeding and Genetics, Cornell University, Ithaca,
New York, USA
Bobby Alton Stewart (125)
Dryland Agriculture Institute, West Texas A&MUniversity, Canyon, Texas, USA
Neil C. Turner (333)
Centre for Legumes inMediterranean Agriculture andUWA Institute of Agriculture,
The University of Western Australia, Crawley, WA, Australia
Zhao-Hui Wang (125)
College of Resources and Environmental Sciences, Northwest Science and
Technology University of Agriculture and Forestry, Yangling, Shaanxi, PR China
PREFACE
Volume 110 contains five excellent reviews dealing with crop and soil
sciences. Chapter 1 is a detailed review on the biogeochemistry, dynamics,
and environmental significance of dissolved organic matter in soils. Chapter 2
deals with prospects and current efforts in using genomic selection in plant
breeding. Chapter 3 is a comprehensive overview on the use of phosphorus
and response to phosphate fertilizers by leguminous and nonleguminous crops.
Chapter 4 deals with the role of mineral nutrition on root growth of crop
plants. Chapter 5 covers the physiology of spikelet development on the rice
panicle and the role that apical dominance plays in grain yield improvement.
I am grateful to the authors for their first-rate contributions.
DONALD L. SPARKS
Newark, Delaware, USA
xi
C H A P T E R O N E
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Dissolved Organic Matter:
Biogeochemistry, Dynamics, and
Environmental Significance in Soils
Nanthi S. Bolan,*,† Domy C. Adriano,‡ Anitha Kunhikrishnan,*,†
Trevor James,§ Richard McDowell,} and Nicola Senesi#
Contents
1. In
s in
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Agronomy, Volume 110 # 2011
-2113, DOI: 10.1016/B978-0-12-385531-2.00001-3 All rig
r Environmental Risk Assessment and Remediation (CERAR), University of Sou
tive Research Centre for Contaminants Assessment and Remediation of the Environ
University of South Australia, Australia
y of Georgia, Savannah River Ecology Laboratory, Drawer E, Aiken, South Carolina
rch, Ruakura Research Centre, Hamilton, New Zealand
rch, Invermay Agricultural Centre, Mosgiel, New Zealand
ent of Agroforestal and Environmental Biology and Chemistry, University of Bari, Ba
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ri,
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2. S
ources, Pools, and Fluxes of Dissolved Organic Matter in Soils
 5
3. P
roperties and Chemical Composition of Dissolved Organic Matter
in Soils
 13
3
.1.
 S
tructural components
 13
3
.2.
 F
ulvic acid—The dominant component
 15
3
.3.
 E
lemental composition
 20
4. M
echanisms Regulating Dynamics of Dissolved Organic
Matter in Soils
 20
4
.1.
 S
orption/complexation
 23
4
.2.
 B
iodegradation
 27
4
.3.
 P
hotodegradation
 28
4
.4.
 L
eaching
 29
5. F
actors Influencing Dynamics of Dissolved Organic Matter in Soils
 30
5
.1.
 V
egetation and land use
 31
5
.2.
 C
ultivation
 32
5
.3.
 S
oil amendments
 33
5
.4.
 S
oil pH
 36
6. E
nvironmental Significance of Dissolved Organic Matter in Soils
 37
6
.1.
 S
oil aggregation and erosion control
 37
6
.2.
 M
obilization and export of nutrients
 38
6
.3.
 B
ioavailability and ecotoxicology of heavy metals
 43
vier Inc.
reserved.
Australia,
t (CRC
SA
Italy
1
2 Nanthi S. Bolan et al.
6
.4.
 T
ransformation and transport of organic contaminants
 50
6
.5.
 G
aseous emission and atmospheric pollution
 58
7. S
ummary and Research Needs
 60
7
.1.
 M
acroscale (landscape to global)
 61
7
.2.
 M
icroscale (water bodies and soil profile)
 61
7
.3.
 M
olecular scale (carbon fractions, organic acids, and
microorganisms)
 61
Ack
nowledgments
 62
Refe
rences
 62
“Dissolved organic matter comprises only a small part of soil organic
matter; nevertheless, it affects many processes in soil and water includ-
ing the most serious environmental problems like soil and water
pollution and global warming.”
(Kalbitz and Kaiser, 2003)
Abstract
Dissolved organic matter (DOM) is defined as the organic matter fraction in
solution that passes through a 0.45 mm filter. Although DOM is ubiquitous in
terrestrial and aquatic ecosystems, it represents only a small proportion of the
total organic matter in soil. However, DOM, being the most mobile and actively
cycling organic matter fraction, influences a spectrum of biogeochemical pro-
cesses in the aquatic and terrestrial environments. Biological fixation of atmo-
spheric CO2 during photosynthesis by higher plants is the primary driver of
global carbon cycle. A major portion of the carbon in organic matter in the
aquatic environment is derived from the transport of carbon produced in the
terrestrial environment. However, much of the terrestrially produced DOM is
consumed by microbes, photo degraded, or adsorbed in soils and sediments as
it passes to the ocean. The majority of DOM in terrestrial and aquatic environ-
ments is ultimately returned to atmosphere as CO2 through microbial respira-
tion, thereby renewing the atmospheric CO2 reserve for photosynthesis.
Dissolved organic matter plays a significant role in influencing the dynamics
and interactions of nutrients and contaminants in soils and microbial functions,
thereby serving as a sensitive indicator of shifts in ecological processes. This
chapter aims to highlight knowledge on the production of DOM in soils under
different management regimes, identify its sources and sinks, and integrate its
dynamics with various soil processes. Understanding the significance of DOM in
soil processes can enhance development of strategies to mitigate DOM-induced
environmental impacts. This review encourages greater interactions between
terrestrial and aquatic biogeochemists and ecologists, which is essential for
unraveling the fundamental biogeochemical processes involved in the synthesis
of DOM in terrestrial ecosystem, its subsequent transport to aquatic ecosystem,
and its role in environmental sustainability, buffering of nutrients and pollutants
(metal(loid)s and organics), and the net effect on the global carbon cycle.
Dissolved Organic Matter 3
1. Introduction
The total organic matter (TOM) in terrestrial and aquatic environ-
ments consists of two operationally defined phases: particulate organic
matter (POM) and dissolved organic matter (DOM). For all practical
purposes, DOM is defined as the organic matter fraction in solution that
passes through a 0.45 mm filter (Thurman, 1985; Zsolnay, 2003). Some
workers have used finer filter paper (i.e., 0.2 mm) in an effort to separate
“true” DOM from colloidal materials, but 0.45 mm filtration appears to be
standard (Buffle et al., 1982; Dafner and Wangersky, 2002). In some litera-
ture, the term dissolved organic carbon (DOC) is used, which represents
total organic carbon in solution that passes through a 0.45 mm filter
(Zsolnay, 2003). Since carbon represents the bulk of the elemental compo-
sition of the organic matter (ca. 67%), DOM is often quantified by its
carbon content and referred to as DOC. In the case of studies involving
soils, the term water-soluble organic matter (WSOM) or water-extractable
organic matter (WEOM) is also used when measuring the fraction of the soil
organicmatter (SOM) extracted with water or dilute salt solution (e.g., 0.5
M K2SO4) that passes through a 0.45 mm filter (Bolan et al., 1996; Herbert
et al., 1993). Recently, the distinction between POM and DOM in the
marine environment is being replaced by the idea of an organic matter
continuum of gel-like polymers, replete with colloids and crisscrossed by
“transparent” polymer strings, sheets, and bundles, from a few to hundreds
of micrometers—referred to as oceanic “dark matter” (Dafner and
Wangersky, 2002).
Dissolved organic matter is ubiquitous in terrestrial and aquatic ecosys-
tems, but represents only a small proportion of the total organic matter in
soil (McGill et al., 1986). However, it is now widely recognized that
because DOM is the most mobile and actively cycling organic matter
fraction, it influences a myriad of biogeochemical processes in aquatic and
terrestrial environments as well as key environmental parameters
(Chantigny, 2003; Kalbitz et al., 2000; McDowell, 2003; Stevenson,
1994; Zsolnay, 2003). Dissolved organic carbon has been identified as one
of the major components responsible for determining the drinking water
quality. For example, DOM leads to the formation of toxic disinfection by-
products (DBPs), such as trihalomethanes, after reacting with disinfectants
(e.g., chlorine) during water treatment. Similarly, DOM can be related to
bacterial proliferation within the drinking water distribution system. There-
fore, the control of DOM has been identified as an important part of the
operation of drinking water plants and distribution systems (Volk et al.,
2002). In aquatic environments, the easily oxidizable compounds in the
DOM can act as chemical and biological oxygen demand compounds,
thereby depleting the oxygen concentration of aquifers and influencing
4 Nanthi S. Bolan et al.
aquatic biota ( Jones, 1992). Dissolved organic carbon can act as a readily
available carbon source for anaerobic soil organisms, thereby inducing the
reduction of nitrate (denitrification) resulting in the release of green house
gases, such as nitrous oxide (N2O) and nitric oxide (NO), which are
implicated in ozone depletion (Siemens et al., 2003). Organic pesticides
added to soil and aquifers are partitioned preferentially onto DOM, which
can act as a vehicle for the movement of pesticide residues to groundwater
(Barriuso et al., 1992). Similarly, the organic acids present in the DOM can
act as chelating agents, thereby enhancing the mobilization of toxic heavy
metals and metalloids [metal(loid)s] (Antoniadis and Alloway, 2002). The
release and retention of DOM are the driving forces controlling a number of
pedological processes including podzolization (Hedges, 1987).
Biological fixation of atmospheric CO2 by higher plants during photo-
synthesis is the primary driver of global carbon cycle. A major portion of the
carbon in aquatic environments is derived from the transport of carbon
produced on land. It has been estimated that worldwide about 210 Mt
DOM and 170 Mt POM are transported annually to oceans from land.
Carbon in the ocean is recognized as one of the three main reservoirs of
organic material on the planet, equal to the carbon stored in terrestrial plants
or soil humus (Hedges, 1987). The terrestrially produced DOM is subject to
microbial- and photodegradation and adsorption by soil and sediments. The
majority of DOM in terrestrial and aquatic environments is returned to the
atmosphere as CO2 through microbial respiration, thereby ultimately
replenishing the atmospheric CO2 reserve for photosynthesis and reinvi-
gorating the global carbon cycle.
Dissolved organic carbon can be envisioned both as a link and bottle-
neck among various ecological compartments. Combined with its dynamic
nature, this enables DOM to serve as a sensitive indicator of shifts in
ecological processes, especially in aquatic systems. Recently, the significance
of DOM in the terrestrial environment has been realized and attempts have
been made to extend this knowledge to DOM dynamics in aquatic envir-
onments. However, DOM dynamics on land are fundamentally different
from those in water, where biomass of primary producers is relatively small,
allochthonous sources of DOM are dominant, the surface area of reactive
solid particles (i.e., sediments) is smaller, and the fate of DOM is strongly
influenced by photolysis and other light-mediated reactions. In contrast, the
dynamics of DOM on land are largely controlled by its interactions with
abiotically and biotically reactive solid components.
Although there have been a number of reviews on the individual
components of DOM in soils (e.g., sources and sink—Kalbitz et al.
(2000); microbial degradation—Marschner and Kalbitz (2003); sorption
by soils—Kaiser et al. (1996)), there has been no comprehensive review
linking the dynamics of DOM to its environmental significance. This
chapter aims to elaborate on the production and degradation of DOM in
Dissolved Organic Matter 5
soils under different landscape conditions, identify its sources and sinks, and
integrate its dynamics with environmental impacts. Understanding the
long-term control on DOM production and flux in soils will be particularly
important in predicting the effects of various environmental changes and
management practices on soil carbon dynamics. Improved knowledge on
the environmental significance of DOM can enhance the development of
strategies to mitigate DOM-induced environmental impacts. It is hoped
that this chapter will encourage greater interaction between terrestrial and
aquatic biogeochemists and ecologists and stimulate the unraveling of
fundamental biogeochemical processes involved in the synthesis and trans-
port of DOM in terrestrial and aquatic ecosystems.
2. Sources, Pools, and Fluxes of Dissolved
Organic Matter in Soils
Nearly all DOM in soils comes from photosynthesis. This represents
the various C pools including recent photosynthates, such as leaf litter,
throughfall and stemflow (in the case of forest ecosystems), root exudates,
and decaying fine roots, as well as decomposition and metabolic by-pro-
ducts and leachates of older, microbiologically processed SOM (Figure 1)
(Guggenberger, et al., 1994a; McDowell, 2003; McDowell, et al., 1998).
The majority of DOM in soils and aquifers originates from the solubilization
of SOM accumulated through vegetation and the addition of biological
waste materials (Guggenberger, et al., 1994b; McDowell, 2003; McDowell,
et al., 1998; Tate and Meyer, 1983). The addition of biological waste
materials, such as poultry and animal manures and sewage sludges, increases
the amount of DOM in soils either by acting as a source of DOM or by
enhancing the solubilization of the SOM.Most biological waste materials of
plant origin contain large amounts of DOM (Table 1) and the addition of
certain organic manures such as poultry manure increases the pH and
thereby enhances the solubilization of SOM (Schindler et al., 1992).
The concentrations of DOM in soils and aquifers are highly susceptible
to changes induced by humans, such as cultivation, fire, clear-cutting,
wetland drainage, acidic precipitation, eutrophication, and climate change
(Kreutzweiser et al., 2008; Laudon et al., 2009; Martinez-Mena et al., 2008;
Mattsson et al., 2009; Yallop and Clutterbuck, 2009). Dissolved organic
matter in environmental samples, such as soils and manures, is often
extracted with water or dilute aqueous salt solutions. Various methods
have been used to measure the concentration of DOM in extracts
(Table 2). These methods are grouped into three categories (Moore,
1985; Sharp et al., 2004; Stewart and Wetzel, 1981; Tue-Ngeun et al.,
2005). The most frequently used method involves the measurement of
B horizon
A horizon
DOM DOM
Litter
layer
Crop
residue
C horizon
Aquifer
Agricultural soilForest soil
11
11
1010
9 9
8 8
6 6
7 7
CO2 CO2
Photosynthesis Photosynthesis
5 5
4 43 32
1
2
Parent/geologic
material
Figure 1 Pathways of inputs and outputs of dissolvedorganic matter (DOM) in forest
and agricultural soils. Inputs: 1, throughfall and stemflow; 2, root exudates; 3, microbial
lysis; 4, humification; 5, litter/and crop residue decomposition; 6, organic amendments;
outputs; 7, microbial degradation; 8, microbial assimilation; 9, lateral flow; 10, sorp-
tion; 11, leaching.
6 Nanthi S. Bolan et al.
absorption of light by the DOM using a spectrophotometer (Stewart and
Wetzel, 1981). The second method involves wet oxidation of samples
containing DOM and the subsequent measurement of the CO2 released
or the amount of oxidant consumed (Ciavatta et al., 1991). This method is
often referred to as chemical oxygen demand (COD). Dichromates or
permanganates are the most common oxidizing agents used in the wet
oxidation of DOM, and the amount of oxidant consumed in the oxidation
of DOM is measured either by titration with a reducing agent or by
calorimetric methods. The third method involves dry oxidation of DOM
to CO2 at high temperature in the presence of a stream of oxygen. The
amount of CO2 produced is measured either by infrared (IR) detector or by
titration after absorbing in an alkali, or by weight gain after absorbing in
ascarite (Bremner and Tabatabai, 1971). The most commonly used dry
combustion techniques include LECOTM combustion and total organic
carbon (TOC) analyzer.
Table 1 Sources of dissolved organic matter input to soils
Sources
Total
organic
matter
(g C kg�1)
Dissolved
organic matter
Reference
(g C
kg�1)
(% of
total
organic
matter)
Pasture leys
Brome grass 13.3 0.041 0.31 Shen et al. (2008)
Clover 15.1 0.039 0.26 Shen et al. (2008)
Crowtoe 10.4 0.036 0.35 Shen et al. (2008)
Lucerne Cv. Longdong 11.4 0.038 0.32 Shen et al. (2008)
Lucerne Cv. Saditi 10.9 0.036 0.33 Shen et al. (2008)
Sainfoin 13.8 0.040 0.29 Shen et al. (2008)
Sweet pea 10.2 0.034 0.33 Shen et al. (2008)
Soil
Forest soil—litter
leachate
60.0 0.026 0.04 Jaffrain et al.
(2007)
Arable soil 12.0 0.150 1.25 Gonet et al. (2008)
Soil under bermuda
grass turf
8.10 0.300 3.70 Provin et al. (2008)
Pasture soil 32.0 1.02 3.18 Bolan et al. (1996)
Pasture soil 82.5 3.12 3.80 Bolan et al. (1996)
Organic amendments
Sewage sludge 420 2.42 0.58 Hanc et al. (2009)
Sewage sludge 321 6.00 1.87 Bolan et al. (1996)
Paper sludge 281 7.19 2.56 Bolan et al. (1996)
Poultry manure 425 8.18 1.92 Bolan et al. (1996)
Poultry littera 377 75.7 20.1 Guo et al. (2009)
Mushroom compost 385 7.10 1.84 Bolan et al. (1996)
Fresh spent mushroom
substrate
288 133 46.2 Marin-Benito et al.
(2009)
Composted spent
mushroom substrate
274 43.4 15.8 Marin-Benito et al.
(2009)
Separated cow manure 456 9.80 2.15 Zmora-Nahuma
et al. (2005)
Poultry manure 425 8.18 1.92 Bolan et al. (1996)
Pig manure 296 6.13 2.07 Bolan et al. (1996)
a Bisulfate amended, phytase-diet Delmarva poultry litter.
Dissolved Organic Matter 7
Plant litter and humus are the most important sources of DOM in soil,
which is confirmed by both field and laboratory (including greenhouse)
studies (Kalbitz et al., 2000; Kalbitz et al., 2007; Muller et al., 2009;
Table 2 Selected references on methods of extraction and analysis of DOM in environmental samples
Samples Extraction of DOM Measurement of DOM Reference
Volcanic ash soils Soil solutions collected by centrifugation of
cores at 7200 rpm; filtration (0.45 mm
filters)
DOC by Shimadzu TOC-
5000 analyzer
Kawahigashi et al.
(2003)
Peat—moorsh soil Soil samples were crushed an passed through
a 1 mm sieve, then heated in a redistilled
water at 100 �C for 2 h under a reflex
condenser; filtration (0.45 mm filters)
DOC by Shimadzu TOC
5050A analyzer
Szajdak et al. (2007)
Soils (medial, amorphic
thermic, Humic
Haploxerands)
Extraction with 0.5 mol L� 1 K2SO4
solution 1:5 (w/v); filtration (Advantec
MFS Nº 5C paper).
TOC by combustion at 675�C
in an analyzer (Shimadzu—
model TOC-V CPN)
Undurraga et al. (2009)
Moss, litter and topsoil
(0–5 cm)
Aqueous samples were estimated for DOC
by oxidation of the sample with a
sulfochromic mixture (4.9 g dm�3
K2Cr2O7 and H2SO4, 1:1, w/w) with
colorimetric detection of the reduced Cr3þ
Colorimeter KFK-3 at 590 nm Prokushkin et al. (2006)
Soil solutions from forested
watersheds of North
Carolina
Samples were filtered through a Whatman
G/F glass fiber filters.
Wet combustion persulfate
digestion followed by
TOC analyzer
Qualls and Haines
(1991)
Organic fertilizer Extracted DOC by 0.01 M CaCl2 solution
with a solid to solution ratio of 1:10 (w/v),
mixed for 30 min at 200 rpm; filtration
(0.45 mm filter)
Shimadzu TOC-5000A
TOC analyzer
Li et al. (2005)
Soil solution and stream waters
along a natural soil catena
Soil solution collected by tension-free
lysimeters
DOC by infrared detection
following persulfate
oxidation
Palmer et al. (2004)
Liquid and solid sludge, farm
slurry, fermented straw, soil,
and drainage water
Water extraction followed by centrifugation
(40,000 � g) and filtration (0.45 mm filter)
Dry combustion (Dhormann
Carbon Analyzer DC-80)
Barriuso et al. (1992)
Soils, peat extract, sludge, pig
and poultry manure and
mushroom compost
Extracted with water (1:3 solid:solution ratio);
centrifugation (12,000 rpm) and filtration
(0.45 mm filter)
Wet chemical oxidation with
dichromate followed by
back titration
Baskaran et al. (1996)
Soil (Entic Haplothord) Extraction with deionized water (1:10 solid:
solution ratio); filtered through 0.45 mm
polysulfore membrane
Dry combustion (TOC
analyzer Shimadzu 5050)
Kaiser et al. (1996)
Pig manure Extracted with water (1:3 solid:solution ratio);
shaken at 200 rpm for 16 h at 4oC;
centrifugation (12,000 rpm) and filtration
(0.45 mm filter)
DOC by Shimadzu TOC-
5000A TOC analyzer
Cheng and Wong
(2006)
Cow manure slurry filtered through 0.45 mm polysulfore
membrane
TOC analyzer using UV
absorbance
Aguilera et al. (2009)
Sewage sludge DOC was extracted in a soil:water ratio of
1:10 (w/v) after 1 h agitation.
Wet combustion with
chromate followed by back
titration
Gascó and Lobo (2007)
River water Natural water from river filtered by
0.22 mm filter
DOC by wet oxidation TOC
analyzer
Krachler et al. (2005)
Peat water Peat water filtered through 0.45 mm
membrane filters
DOC was analyzed using a
high-temperature catalytic
oxidation method
(Dohrman DC-190
analyzer)
Rixen et al. (2008)
River water Filtered through 0.7 mm glass fiber filter In situ optical technology
using fluorescence
Spencer et al. (2007)
(continued)
Table 2 (continued)
Samples Extraction of DOM Measurement of DOM Reference
Sea water Filtered through 0.45 mm polysulfore
membrane
High-temperature
combustion instrument to
measure isotope
composition of DOC
Lang et al. (2007)
Freshwater Filtered through 0.7 mm glass fiber filter Acid-peroxydisulfate
digestion and high-
temperature catalytic
oxidation (HTCO) with
UV detection
Tue-Ngeun et al. (2005)
Effluent water – In situ UV spectrophotometer Rieger et al. (2004)
Groundwater, lake water,
and effluent
– High-performance liquid
chromatography-size
exclusion chromatography-
UVA fluorescence system
Her et al. (2003)
Sea water and effluent Filtered through 0.7 mm glass fiber filter Measurement of carbon
atomic emission intensity in
inductively coupled plasma
atomic emission
spectrometry (ICP-OES)
Maestre et al. (2003)
Lake water Water samples filtered using precombusted
GF/F filters
TOC analyzer (TOC 5000;
Shimadzu)
Ishikawa et al. (2006)
Soil solution and stream water
from forested catchments
Samples were filtered through
0.45 mm filters
DOC by Shimadzu TOC
5050A analyzer
Vestin et al. (2008)
Dissolved Organic Matter 11
Sanderman et al., 2008). In forest ecosystems, which are the most intensively
studied with regard to C cycling and its associated DOM dynamics, the
canopy and forest floor layers are the primary sources of DOM (Kaiser et al.,
1996; Kalbitz et al., 2007; Park and Matzner, 2003). However, it is still
unclear whetherDOM originates primarily from recently deposited litter or
from relatively stable organic matter in the deeper part of the organic
horizon (Kalbitz et al., 2007).
In a temperate, deciduous forest, the source of DOM leaching from the
forest floor (O layer) is generally a water-soluble material from freshly fallen
leaf litter and throughfall (Kalbitz et al., 2007; Qualls et al., 1991). Appar-
ently all of the DOM and dissolved organic N (DON) could have origi-
nated from the Oi (freshly fallen litter) and Oe (partially decomposed litter)
horizons. They further observed that, while about 27% of the freshly shed
litter C was soluble, only 18.4% of the C input in litterfall was leached in
solutions from the bottom of the forest floor. Virtually all the DOM leached
from the forest floor appeared to have originated from the upper forest
floor, with none coming from the lower forest floor—an indication of the
role of this litter layer as a sink. The role of freshly deposited litter as DOM
source was further corroborated by laboratory studies (Magill and Aber,
2000; Moore and Dalva, 2001; Muller et al., 2009; Sanderman et al., 2008).
Michalzik andMatzner (1999) found high fluxes of DOM from the Oi layer
than from the Oe and Oa layers and indicated that the bottom organic layers
acted instead as a sink rather than as a source of DOM. Logically, however,
because of the more advanced state of decomposition, the bottom litter
layers could produce more DOM than the surface layer. Indeed, Solinger
et al. (2001) measured greater DOM fluxes out of the Oa than out of the Oi
layer. Recently, Froberg et al. (2003) and Uselman et al. (2007) confirmed
with 14C data that the Oi layer is not a major source of DOM leached from
the Oe layer.
In a comprehensive synthesis of 42 case studies in temperate forests,
Michalzik et al. (2001) observed that, although concentrations and fluxes
differed widely among sites, the greatest concentrations of DOM (and
DON) were generally observed in forest floor leachates from the A horizon
and were heavily influenced by annual precipitation. However, somewhat
surprisingly, there were no meaningful differences in DOM concentrations
and fluxes in forest floor leachates between coniferous and hardwood sites.
The flux of soluble organic compounds from throughfall and the litter layer
could amount to 1–19% of the total litterfall C flux and 1–5% of the net
primary productivity (Froberg et al., 2007; McDowell and Likens, 1988;
Qualls et al., 1991). Nearly one-third of the DOM leaving the bottom of the
forest floor originated from throughfall and stemflow (Qualls et al., 1991;
Uselman et al., 2007). Values for the potential solubility of litter in the field
and in laboratory studies are in the 5–25% range of the litter dry mass and
5–15% of the litter C content (Hagedorn and Machwitz, 2007; McDowell
12 Nanthi S. Bolan et al.
and Likens, 1988; Muller et al., 2009; Sanderman et al., 2008; Zsolnay and
Steindl, 1991).
In typical soils, DOM concentrations may decrease by 50–90% from the
surface organic layers to mineral subsoils (Cronan and Aiken, 1985; Dosskey
and Bertsch, 1997; Worrall and Burt, 2007). Similarly, fluxes of DOM in
surface soil range from 10 to 85 g C m�2 yr�1, decreasing to 2–40 g C m�2
yr�1 in the subsoils (Neff and Asner, 2001).
In cultivated and pastoral soils, plant residues provide the major source of
DOM, while in forest soils, litter and throughfall serve as the major source
(Ghani et al., 2007; Laik et al., 2009). In forest soils, DOM represents a
significant proportion of the total C budget. For example, Liu et al. (2002)
calculated the total C budgets of Ontario’s forest ecosystems (excluding peat
lands) to be 12.65 Pg (1015g), including 1.70 Pg in living biomass and 10.95
Pg in DOM in soils. Koprivnjak and Moore (1992) determined DOM
concentrations and fluxes in a small subarctic catchment, which is composed
of an upland component with forest over mineral soils and peat land in the
lower section. DOM concentrations were low (1–2 mg L�1) in precipita-
tion and increased in tree and shrub throughfall (17–150 mg L�1), the
leachate of the surface lichens and mosses (30 mg L�1), and the soil A
horizon (40 mg L�1). Concentrations decreased in the B horizon (17 mg
L�1) and there was evidence of strong DOM adsorption by the subsoils.
Khomutova et al. (2000) examined the production of organic matter in
undisturbed soil monoliths of a deciduous forest, a pine plantation, and a
pasture under constant temperature (20 �C) and moisture. After 20 weeks of
leaching with synthetic rain water at pH 5, the cumulative values of DOM
production followed: coniferous forest > deciduous forest > pasture, the
difference being attributed to the nature of carbon compounds in the
original residues. The residues from the coniferous forest were found to
contain more labile organic components.
Among ecosystems types, Zsolnay (1996) indicated that DOM tends to
be greater in forest than agricultural soils: 5–440 mg L�1 from the forest
floor compared with 0–70 mg L�1 from arable soils. Other studies have also
indicated greater concentrations of DOM and concentrations in grasslands
than in arable soils (Ghani et al., 2007; Gregorich et al., 2000; Haynes,
2000). In general, DOM concentration decreases in the order: forest floor>
grassland A horizon > arable A horizon (Chantigny, 2003).
The rhizosphere is commonly associated with large C flux due to root
decay and exudation (Muller et al., 2009; Uselman et al., 2007; Vogt et al.,
1983). Microbial activity in the rhizosphere is enhanced by readily available
organic substances that serve as an energy source for these organisms
(Paterson et al., 2007; Phillips et al., 2008). Because of their turnover, soil
microbial biomass is also considered as an important source of DOM in soils
(Ghani et al., 2007; Steenwerth and Belina, 2008; Williams and Edwards,
1993). Thus, microbial metabolites may represent a substantial proportion
Dissolved Organic Matter 13
of the soil’s DOM. It may well be that the rate of DOM production and
extent of DOM dynamics in soil is regulated by the rate of litter/residue
incorporation in soils, kinetics of their decomposition, and various biotic
and abiotic factors (Ghani et al., 2007; Kalbitz et al., 2000; Michalzik and
Matzner, 1999; Zech et al., 1996).
In summary, the various C pools in an ecosystem represent the sources of
DOM in soils. Due to their abundance, recently deposited litter and humus
are considered the two most important sources of DOM in forest soils.
Similarly, recently deposited crop residues and application of organic
amendment such as biosolids and manures are the most important sources
of DOM in arable soils. However, the role of root decay and/or exudates
and microbial metabolites cannot be downplayed in both forested and arable
ecosystems.
3. Properties and Chemical Composition of
Dissolved Organic Matter in Soils
3.1. Structural components
Because DOM is a heterogeneous composite of soluble organic compounds
arising from the decomposition of various carbonaceous materials of plant
origin, including soluble microbial metabolites from the organic layers in
the case of forest ecosystem, DOM constituents can be grouped into
“labile” DOM and “recalcitrant” DOM (Marschner and Kalbitz, 2003).
Labile DOM consists mainly of simple carbohydrate compounds (i.e.,
glucose and fructose), low molecular weight (LMW) organic acids, amino
sugars, and LMW proteins (Guggenberger et al., 1994b; Kaiser et al., 2001;
Qualls and Haines, 1992). Recalcitrant DOM consists of polysaccharides
(i.e., breakdown products of cellulose and hemicellulose) and other plant
compounds, and/or microbially derived degradation products (Marschner
and Kalbitz, 2003) (Table 3). Soil solution DOM consists of LMW carbox-
ylic acids, amino acids, carbohydrates, and fulvic acids—the first comprising
less than 10% of total DOM in most soil solutions and the last (i.e., fulvic
acid) being typically the most abundant fractions of DOM (Strobel etal.,
1999, 2001; Thurman, 1985; van Hees et al., 1996).
Dissolved organic matter is separated into fractions based on solubility,
molecular weight, and sorption chromatography. Fractionation of DOM by
molecular size and sorption chromatography separate DOM according to
properties (hydrophobic and hydrophilic) which regulate its interaction
with organic contaminants and soil surfaces. The most common technique
for the fractionation of aquatic DOM is based on its sorption to non-ionic
and ion-exchange resins (Leenheer, 1981).
Table 3 Components identified in specific fractions of dissolved organic matter
Fraction Compounds Reference
Hydrophobic neutrals Hydrocarbons Polubesova et al. (2008)
Chlorophyll Albrechtova et al. (2008)
Carotenoids Leavitt et al. (1999)
Phospholipids Yoshimura et al. (2009)
Weak (phenolic) hydrophobic acids Tannins Suominen et al. (2003)
Flavonoids
Other polyphenols Hernes et al. (2007)
Vanillin Suominen et al. (2003)
Strong (carboxylic) hydrophobic acids Fulvic acid and humic acid Christensen et al. (1998)
Humic-bound amino acids and peptides Lytle and Perdue (1981)
Humic-bound carbohydrates Volk et al. (1997)
Aromatic acids (including phenolic
carboxylic acids)
Gigliotti et al. (2002)
Oxidized polyphenols Serrano (1994)
Long-chain fatty acids Jandl et al. (2002)
Hydrophilic acids Humic-like substances with lower molecular
size and higher COOH/C ratios
Oxidized carbohydrates with COOH groups
Small carboxylic acids Obernosterer et al. (1999)
Inositol and sugar phosphates Monbet et al. (2009)
Hydrophilic neutrals Simple neutral sugars Borch and Kirchman (1997)
Non-humic-bound polysaccharides Rosenstock et al. (2005)
Alcohols Chefetz et al. (1998)
Bases Proteins Schulze (2004)
Free amino acids and peptides Yamashita and Tanoue (2004)
Aromatic amines Yamashita and Tanoue (2004)
Amino-sugar polymers (such as from microbial cell walls) Jones et al. (2005)
Dissolved Organic Matter 15
Bolan et al. (1996) examined the distribution of various molecular
weight fractions in the DOM extracts of various sources including soil,
manures, composts, and sewage sludge. The DOM samples varied in the
relative distribution of molecular weight fractions. The DOM from sewage
sludge and poultry manure has a greater proportion of DOM in LMW
fractions than DOM from soil or stream water. These results are consistent
with the results for chemical oxidation, indicating that LMW fractions are
more readily oxidized than the high molecular weight fractions.
Dai et al. (1996) examined the structural composition and fractions
(hydrophobic and hydrophilic acids and hydrophilic neutrals) of DOM
from forest floor leachates over a 2-year period using 13C NMR spectros-
copy. Total DOM in forest floor leachates ranged from 7.8 to 13.8 mmol
L�1 with an average of 8.6 mmol L�1. These solutions were enriched with
organic acids that averaged 92% of the total DOM. The 13C NMR data
suggested that alkyl, carbohydrate, aromatic, and carboxylic C were the
primary constituents of DOM fractions. Compositional changes of C with
depth were observed, aromatic and carbohydrate decreased, whereas alkyl,
methoxy, and carbonyl moieties increased with depth. Hydrophobic acids
contained high contents of aromatic C, whereas hydrophilic acids primarily
comprised carboxylic C. Hydrophilic neutrals were rich in carbohydrate C.
Engelhaupt and Bianchi (2001) noticed that DOM from soils and leaf
litter was dominated by aliphatic (41%), carbohydrate (33%), and carboxyl
(16%) carbon, with relatively low aromatic carbon (10%). This study
demonstrated that lignin and other compounds from terrestrially derived
organic matter in sediments and adjacent soils were not a significant source
of soluble moieties that enter the HMW DOM pool of tidal streams.
Maurice et al. (2002) observed that the contribution of soil pore water
relative to groundwater controlled not only the concentration, but also the
average physicochemical characteristics of the DOM in streams. Combined
field and laboratory experiments suggested that preferential adsorption of
HMW and aromatic DOM components to mineral surfaces within the
lower soil horizons resulted in more aliphatic groundwater DOM pool.
Low flow periods resulted in an aliphatic dominated DOM in streams,
whereas higher flow periods resulted in more aromatic downstream surface
water DOM pool.
3.2. Fulvic acid—The dominant component
The fractions of soil humic substances that are water soluble at any pH
above 1–2, that is, fulvic acids (FAs), are very abundant and important
components of soil DOM and a large number of studies have focused on
the structure and chemical composition of FAs in DOM (Plaza and Senesi,
2009; Senesi and Loffredo, 1999; Senesi and Plaza, 2007). The FAs feature
composition, structure, and chemical and biochemical properties that
16 Nanthi S. Bolan et al.
definitely distinguish them from the other typical components of DOM, all
of which belong to definite organic chemical classes. On the contrary, soil
FAs are not defined by a unique chemical formula and do not belong to any
of the known chemical classes of organic compounds. The FAs consist of a
physically and chemically heterogeneous mixture of relatively low molecu-
lar weight (500–2000 Da), yellow-to-light-brown/reddish organic mole-
cules of mixed aliphatic and aromatic nature, and bearing acidic functional
groups (mainly carboxylic and phenolic OH), which are formed by second-
ary synthesis reactions of recalcitrant compounds with products of microbial
and chemical decay and transformation of biomolecules originated from
organisms during life and after death (Senesi and Loffredo, 1999). These
distinctive features confer to the FA fraction of soil DOM unique behavior
and performances in soil chemical and biological reactivity, especially
toward metal(loid) ions and organic contaminants.
The major oxygen-containing functional groups in FA are COOH and
phenolic OH groups, whereas alcoholic OH and carbonyl and methoxyl
groups are found in smaller amounts. During humification, COOH and
carbonyl groups have been found to increase, whereas phenolic and alco-
holic OH and methoxyl groups decreased.
FAs behave like weak acid polyelectrolytes whose acidic properties have
been studied by base titration using potentiometric, conductometric, high-
frequency, and thermometric techniques. Such occurrence has been found
in FAs of a continuous and complex spectrum of nonidentical acidic
functional groups with pKa values that span a very wide range as a function
of FA concentration and presence of neutral salts. Although there is dis-
agreement about the pKa values of soil FAs that are recorded in the litera-
ture, the pKa provides a convenient means of comparing the strengths of
acidic groups in FAs from various sources and for any given FA as affected
by neutral salts and an indication of the expected degree of ionization at
various pHs. Several mathematical models have been applied to describe
proton binding by FAs including continuous distribution models and affin-
ity spectrum models.
FAs are a variable-charge soil component with a low-point-of-zero
charge of about 3. Thus, FAs are negatively charged at pH >3, and
COOH and phenolic OH groups of FAs are among the major contributors
to the negative charge of soil. In general, the cation-exchange capacity of
FAs increases with increasing the degree of humification and soil pH.
The molecular weight (MW), size, and shape are very important basic
properties of FA. However, several problems have been encountered when
dealing with the measurement of these properties that are greatly dependent
on the physical state and concentration of the FA, and pH and ionic strength
of the medium. FAs are polydisperse materials, that is, they exhibit a range
of MW that may vary from a few hundred to a couple of thousand Daltons.
Because of the polydispersed nature of FAs, methods that could provide
Dissolved Organic Matter 17
distribution patterns of MW for FA have beenapplied. Furthermore, the
average MW of polydispersed systems can be expressed in several ways
depending on the physical method of determination. These include the
number-average MW, the weight-average MW, the z-average MW, and
viscosity-average MW. The weight-average MW is generally considered
the most representative average MW value because it better correlates with
the molecular properties of FA.
The sizes and shapes of FAs, that is, their morphological conformation, can
be directly observed by the use of transmission and scanning electron micro-
scopes (TEM and SEM). However, sample preparation, especially drying
procedure, was found to affect markedly the morphological features of FAs
and thus represent themost critical aspect of electronmicroscopy application to
the study of FAs. Furthermore, the pH of the medium and FA concentration
were found to be crucial for determining the conformation of FAs. In particu-
lar, at acidic to neutral pH (from 2 to 7), FA exhibited the shape of elongated,
linear, or curved fibers that tended to become thinner with increasing pH, and
of bundles of fibers that tended to become predominant at pH 6, and to give a
fine network at pH 7. At higher pH (8 and 9), the FA assumed a sheet-like
structure of increasing thickness, whereas at pH10, a fine homogeneous gram-
like shape was apparent. At low concentrations, the FA particles assumed an
almost spheroidal shape with tendencies to coalesce to round-shaped aggre-
gates or linear, chain-like shapes. At intermediate concentrations fiber-like
shapeswere formedbyFA,whereas at the highest concentrations parallel arrays
of filaments tended to coalesce to sheet-like shapes.
A number of chemical and structural information on FAs could be
provided by the use of chemical and thermal degradation methods including
hydrolysis; reduction with sodium amalgam and by zinc dust distillation and
fusion; oxidation with alkaline permanganate, alkaline cupric oxide, and
peracetic acid; degradation with sodium sulfide and phenol; thermogravi-
metry; differential thermal analysis; and differential thermal calorimetry
(e.g., Chen et al., 1978b). However, the most modern and powerful
pyrolysis techniques have provided the most interesting results. Pyrolysates
of FA contain a rich mixture of products in various proportions that can be
related to their constituent building blocks, lateral chains, and functional
groups. These include high levels of polysaccharides, phenolic constituents,
n-alkanes, fatty acids, diols, sterols, alkyl mono- and di-esters, and furan
rings; low levels of polypeptide products, lignin products, microbially
synthesized polyphenols, and aromatic hydrocarbons; and various levels of
substituted polycarboxylic acids, amino sugars, lipids, and other aliphatic
constituents. Advanced methods of analytical pyrolysis, especially Curie-
point pyrolysis–gas chromatography–mass spectrometry and pyrolysis–
FIMS, made possible the identification of chemical building blocks in FA
and provided a molecular chemical basis for modeling a structural network
for FAs in which aromatic rings are joined by alkyl chains.
18 Nanthi S. Bolan et al.
Spectroscopic techniques, such as infrared (IR), nuclear magnetic reso-
nance (NMR), fluorescence, and electron spin resonance (ESR) spectro-
scopies, have had wide applications to the study of FAs that have enhanced
our knowledge of their chemical structure and properties (Senesi, 1990a,b,
c; Senesi et al., 1989). IR spectroscopy has been the most used classical
spectroscopic technique in the study of FAs and has allowed the qualitative
and semiquantitative identification of several typical components present at
various levels in the structure of FAs. These include short- and long-chain
aliphatic CH bonds, COOH and other carbonyl groups, variously substi-
tuted aromatic structures, amide groups, nonaromatic double bonds, con-
jugated ketones and quinones, phenolic and alcoholic groups, aryl ethers,
and polysaccharides.
The rapidly advancing powerful NMR techniques are among the most
useful tools currently available for the qualitative and quantitative study of
structural components and functional groups of FAs. The 1H NMR has
allowed the identification of several hydrogenated components (protons)
present in FAs. These include terminal methyl and methylene groups,
methyl and methylene groups bound to alicyclic or aromatic rings, olefins,
phenols, and COOH groups. The dominant peak area of the CP MAS 13C
NMR spectra of FAs is the C–O chemical shift region primarily due to
polysaccharides. Other well-resolved peaks are assigned to (a) unsubstituted
aliphatic C comprising methyl, methylene, and methine groups; (b) C in C–
O of methoxyl groups; (c) C in all other aliphatic and C–O and C–N
groups; (d) anomeric C; (e) aromatic C; and (f) carbonyl C in carboxyl,
ester, and amide groups. Additional two poorly resolved peaks are assigned
to (a) aromatic C in phenolic groups, aromatic amine groups, and aromatic
ethers and (b) carbonyl C in ketonic groups. For quantitative analysis, peak
areas of the spectrum corresponding to the various chemical shift zones are
often measured by integration, thus providing the distribution of various C
types in FA. 15N, 31P, and other nuclei NMR has also been applied in FA
studies with various success.
Fluorescence monodimensional spectroscopy in the emission, excita-
tion, and synchronous scan modes and bi- and tri-dimensional fluorescence
spectroscopies have also been widely applied in the study of FAs. The
fluorescence emission spectra of FAs generally consist of a unique broad-
band with a maximum wavelength that ranges from 445 to 465 nm.
Fluorescence excitation spectra of FAs generally feature one main peak in
the intermediate region of the spectrum (around 390 nm) with additional
minor peaks and shoulders at longer and/or shorter wavelengths. FAs
generally exhibit fluorescence synchronous scan spectra that are more
structured than the corresponding emission and excitation ones, featuring
two main peaks at long (450–460 nm) and intermediate (390–400 nm)
wavelengths, often with some less intense peaks and/or shoulders at both
sides (Miano and Senesi, 1992). Bi- and tri-dimensional fluorescence has
Dissolved Organic Matter 19
also been applied with success in the study of FAs. Analysis of fluorescence
spectra has provided some useful and unique information on the structure
and functionalities of FAs. For example, hydroxyl- and methoxy-coumarin-
like structures, such as esculetin and scopoletin, originated from lignin,
chromone, and xanthone derivatives; Schiff-base fluorophores derived
from polycondensation reactions of carbonyls with amino groups; benzene
rings bearing an hydroxyl conjugated to a carbonyl, methylsalicylate moi-
eties, and dihydroxybenzoic acid units such as protocatechuic, caffeic, and
ferulic acids have all been suggested as possibly responsible for fluorescence
of FAs at various wavelengths.
Fluorescence properties and intensity of FAs have been shown to be
extensively affected by some molecular parameters and conditions of the
medium. These include origin and nature, molecular weight and concen-
tration of FA in solution, and pH and ionic strength of the medium.
Application of ESR spectroscopy has provided important information
on the existence, nature, and concentration in FAs of indigenous organic-
free radicals and complexed paramagnetic metal(loid) ions such as Cu, Fe,
Mn, and V, which may be involved at various stages in several important
chemical, biochemical, and photochemical processes occurring in soil and
water systems. ESR data are consistent with the existence in FAs of indige-
nous semiquinone radical units extensively conjugated to aromatic rings.
The concentration of organic free radicals (between about 1016 and 1018
spins g�1) is probably the most important datum that can be obtained from
the ESR spectrum of FA and has been shown to depend on numerous
measurement conditions and environmental factors (Senesi et al., 1977a,b).A marked increase in free radical concentration of FAs is caused by raising
the pH or temperature, chemical reduction, UV–Vis light irradiation, and
acid hydrolysis. However, the increase was shown not to be sustained in
time but followed by a gradual decrease soon after the maximum value was
attained. On the contrary, mild chemical or electrochemical oxidation,
methylation, and an increase in neutral electrolyte concentration often
produced a time- and pH-dependent decrease of free radical concentration
in FA. The 10-fold decrease of free radical concentration measured for some
FAs confirmed that phenolic OH groups are the most important electron
donors responsible for the formation and existence of free radicals in FA.
The effect of oxidation could be reversed, however, by treatment with a
reductant or by light irradiation of the FA sample. The accumulated ESR
evidence supports the existence of a quinone–hydroquinone electron
donor–acceptor (or charge transfer) system for the reversible generation
and maintenance of free radicals of semiquinonic nature in FAs. Two classes
of free radicals of similar nature, but of different stability, were suggested to
exist in FAs. Besides indigenous or “native” semiquinone radicals, which
are stable over long time spans and survive in any conditions of the system,
“transient” or short-lived semiquinone radicals can be generated by reaction
20 Nanthi S. Bolan et al.
of quinone and hydroquinone moieties in FA, which can only persist in
relatively short time spans. The free radical concentration in FAs was also
shown to be directly related to their color, degree of aromaticity, and
molecular size and complexity.
3.3. Elemental composition
The elemental composition of DOM depends on its origin (Table 4). The
major elements accompanying carbon include oxygen, hydrogen, nitrogen,
phosphorus, sulfur, and trace amounts of various cations including calcium,
potassium, magnesium, and metal(loid)s including aluminum, iron, zinc,
and copper. For example, Kaiser (2001) found that the organic forest floor
layers were large sources for DOC, DON, DOP, and DOS. The dissolved
organic nutrients were mainly concentrated in the hydrophilic DOM
fraction, which proved to be more mobile in mineral soil pore water than
the hydrophobic one. Consequently, the concentrations and fluxes of
dissolved organic nutrients (DON, DOP, and DOS) decreased less with
depth than those of DOC.
The average elemental composition (in percentage) of soil FA is C, 45.7;
O, 44.8; H, 5,4, N, 2.1; S, 1.9 (Senesi and Loffredo, 1999). However, the
composition range of FAs varies at some extent as a function of several
factors including climate, parent material, vegetation, soil age, and pH
(Chen et al., 1978a; Senesi et al., 1989). Methods used for soil FA extraction
may also affect the analytical results and may cause lack of reproducibility.
Typical O/C and H/C ratios of soil FAs are 0.7 and 1.4, respectively. High
O/C ratios reflect high amounts of oxygenated functional groups such as
COOH and carbohydrates, whereas low H/C ratios would indicate a high
contribution of aliphatic components in FA.
The elemental composition of DOM in relation to mobilization of
nutrients is discussed in Section 6. Thus, the chemical composition and
structural properties of various components in DOM are influenced by
sources and their decomposition stage and play a vital role in the interactions
of DOM with heavy metal(loid)s, nutrients, and pesticides.
4. Mechanisms Regulating Dynamics of
Dissolved Organic Matter in Soils
The net pool of DOM in soils is the result of various biogeochemical
processes, resulting in a balance between the input and output of organic C
in the forest floor (or surface soils in arable and grassland soils). These
biological (biodegradation/decomposition, biotransformation), chemical
(sorption, complexation, photodegradation), and physical (leaching,
Table 4 Elemental composition of dissolved organic matter
Source
Elemental composition (mg L�1)
Measured in ReferenceCarbon Nitrogen Phosphorus Sulfur Metals
Pasture soil 28.8 Soil extract Stumpe and Marschner (2010)
Arable soil Soil extract
Forest soil 7.9–13.9 0.9–1.2 Soil solution Möller et al. (2005)
Rhizosphere soil
(Grassland)
11–32 2.5–9 Soil solution Khalid et al. (2007)
Grassland soil 2.5–10 Soil solution Jones et al.(2004)
Grassland soil 0.017–0.133 Soil solution McDowell (2005)
Wetland soil 5–140 0.03–2.4 Soil solution D’Amore et al. (2009)
Cattle manure 1807.2 Stumpe and Marschner (2010)
Forest floor 0.253 Seepage water Kaiser and Guggenberger (2005b)
Forest floor-derived
from litter
45,1000 17,000 1000 3400 Soil solution Kaiser (2001)
Forest floor 23 1.18 0.06 0.25 Soil solution Kaiser and Guggenberger (2005a)
Sewage sludge-
amended soil
61.7 Cd-0.13
Ni-271.42
Zn-145.31
Sludge soil
solution
Antoniadis et al. (2007)
Sewage sludge 277.7 Liquid sewage
sludge
Zhaohai et al. (2008)
Sewage sludge 4395 Cu-0.905
Ni-2.215
Zn-2.315
Sludge solution Ashworth and Alloway (2004)
Stream 11–46 0.2–0.6 Stream water D’Amore et al. (2009)
(continued)
Table 4 (continued)
Source
Elemental composition (mg L�1)
Measured in ReferenceCarbon Nitrogen Phosphorus Sulfur Metals
Surface water 38.2 Cu-0.009
Pb-0.018
Zn-0.371
Cd-0.0004
Stream water Karlik and Szpakowska (2001)
Groundwater 10.5 Cu-0.012
Pb-0.029
Zn-0.506
Cd-0.0009
Stream water Karlik and Szpakowska (2001)
Poultry litter 16,600 2160 Poultry litter
extract
Goyne et al. (2008)
Dissolved Organic Matter 23
eluviation) processes are in turn moderated by biotic and abiotic factors that
include soil pH, organic carbon and clay contents, microbial activity, and
environmental factors including temperature and moisture content
(Table 5). The role of these factors in controlling the dynamics of DOM
is discussed in Section 5.
4.1. Sorption/complexation
Like any other solute in soils, DOM undergoes both sorption and complex-
ation reactions (Guggenberger and Kaiser, 2003; Kothawala et al., 2009;
Remington et al., 2007; Vandenbruwane et al., 2007; Yurova et al., 2008).
While sorption results in the retention of DOM with soil components and
subsequent retardation of its mobility and degradation, complexation can
result in the formation of both soluble and insoluble DOM–metal(loid)
complexes, thereby affecting both movement and degradation. While solu-
ble DOM–metal(loid) complexes enhance the movement of DOM in soils,
insoluble complexes result in the retardation of DOM movement
(Guggenberger and Kaiser, 2003; Jansen et al., 2005; Martin and
Goldblatt, 2007). Complexation of DOM with metal(loid) ions controlling
the movement and bioavailability of both DOM and metal(loid)s is
Table 5 Mechanisms and factors regulating the dynamics of dom in different land
use systemsa
Mechanism/Factor
Agricultural lands Forest lands Wetlands
Arable Pasture/Prairie Upland Savannah Rice Swamps
Sorption xx xx xx xx xx x
Complexation x xx x xx xx xxx
Bidegradation xx xx xx xx xx xx
Biotransformationb NA NA NA xx xxx xxx
Photodegradation x x x x xx xxx
Leaching xx xx xx xx xxx xxx
Vegetation xx xx xx xx xx x
Cultivation xxx x NA NA NA NA
Soil amendments xxx xx x NA xxx NA
Soil pH xx xx xx x xx x
Clay mineralogy xx xx xx x xx x
Metal oxides xx xx xx x xx x
Organic matter xx xxx xxx xxx xx x
a Degree of importance: x, low; xx, medium; xxx, high; NA, not applicable.
b This Mechanism refers primarily to methane formation in reducing conditions such as rice paddy,
swamps and to some extent savannahs.
24 Nanthi S. Bolan et al.
discussed in detail in Section 6.3. Briefly, when DOM percolates in the soil
profile, it may interact with metal oxide surfaces, thereby forming a “shield”
against microbial attack. In acid forest soils, Al and Fe can form relatively
stable complexes with DOM, which can enhance solubility and transport, as
might be the case during podzolization (Blaser, 1994; Jansen et al., 2005).
However, complexation of potentially toxic metal(loid)s maynot result in
diminished biodegradability of DOM, but may even enhance microbial
activity by sequestering the toxic effects from free metal(loid) ion activity
(Apte et al., 2005; Marschner and Kalbitz, 2003). Similarly, formation of
stable complexes between DOM and certain heavy metal(loid)s ions such as
Cu, Hg, and Pb can alter the metal(loid) toxicity to fish and other aquatic
organisms (Adriano, 2001; Alberts et al., 2001; Martin and Goldblatt, 2007).
In temperate soils, the greatest concentrations of organicC typically occur
in the organic layers and the mineral topsoil (A) horizon. However, based on
total soil mass in the various horizon depths in the soil profile, subsoil (B and
C) horizons could account for greater amounts of organic C ( Johnson et al.,
2009; Paul et al., 2002; Schulze et al., 2009; Ziegler, 1991). In investigating
two temperate acid forest soils, Kaiser et al. (2002) observed that the organic
forest floor layer and B andC horizons contained 40–50% of the total DOM.
The ultimate fate of DOM in the soil profile is largely influenced by the
nature and extent of soil mineral—organic carbon—microbe interactions
(Huang et al., 2005a; Young et al., 2008). In essence, partitioning of the
DOM between the aqueous (i.e., soil solution) and the solid phase (i.e., soil
matrix) is controlled by the properties and composition of DOM, microbial
population, and mineralogical and chemical properties of the soil (Adriano,
2001; Guggenberger and Kaiser, 2003; Kothawala et al., 2009; Stevenson,
1994). For example, clays may interact directly with microbes, thereby
altering the rate and pathways of microbial metabolism; modify the aqueous
phase environment (e.g., buffering the pH that affects microbial and enzyme
activity and chemical speciation of contaminant chemicals); and bind extra-
cellular enzymes altering their activity (Grandy et al., 2008; Huang et al.,
2005b; Sollins et al., 1996).
Clay is a generic term that includes layer and amorphous aluminosilicates
and the sesquioxides (i.e., oxides, hydroxides, and oxyhydroxides of Al and
Fe) that provide the majority of surface area for the sorption of DOM and
other solutes in soil. Organic–mineral interactions range in strength from
strong ligand exchange to weaker anion-exchange reactions (McBride,
1994). The bonding mechanisms of DOM onto the soil solid phase have
already been elucidated by Gu et al. (1994) and Sollins et al. (1996). This
includes bonding of negatively charged organics by ligand exchange espe-
cially in oxide-rich and allophanic soils, positively charged organics into
negative surfaces by cation exchange, anion exchange onto subsoils and
variably charged soils, and the less important mechanisms such as cation
bridging, water bridging, hydrogen bonding, and van der Waals forces.
Dissolved Organic Matter 25
The affinity of soils for DOM is influenced by several properties.
Correlations between the extent of partitioning of DOM and surface area
of clay, organic C, dithionite–citrate–bicarbonate-extractable Fe, and oxa-
late-extractable Fe and Al have been reported (Donald et al., 1993;
Guggenberger and Kaiser, 2003; Kaiser et al., 1996; Kothawala et al.,
2009; Nelson et al., 1993). Tipping (1981) reported that the surface area is
the main factor influencing DOM sorption to Fe oxides/hydroxides.
Donald et al. (1993) measured the sorption of DOM and its fractions by
soil horizons from a catenary sequence. Variation in DOM sorption among
the soil horizons was related to differences in the clay content and citrate–
dithionate-extractable Fe, Al, and Mn. The hydrophobic acid and the
hydrophilic acid fractions were the most abundant in the soil solution
(72% of the total DOM) and accounted for most of the sorption of DOM
in the Bt and C horizons.
Moore et al. (1992) obtained DOM sorption isotherms for 48 soil
samples derived from Humaquepts, Inceptisols, and Spodosols in southern
Quebec using a DOM solution derived from a swamp peat. Forty-six
samples had DOM sorption adequately represented by the linear initial
mass isotherm. Null-point DOM concentrations (DOMnp), where there
is zero net DOM sorption, ranged from 6.7 to 85.4 mg L�1. Distribution
coefficients (kd) averaged 1.00 � 10�2 m3 kg�1, suggesting that DOM
sorption by soils is of moderate strength compared with inorganic anions.
DOMnp values were positively correlated to organic C content and nega-
tively correlated to oxalate-extractable Al and dithionite-extractable Fe,
which explained 70% of the variation in DOMnp. Recently, Kothawala
et al. (2009) noticed that poorly crystalline Al oxides exerted a stronger
influence than Fe oxides on maximum sorption capacity of DOM for 52
mineral soil samples from 17 temperate and boreal soil profiles.
Kaiser and Zech (1997) obtained DOM sorption isotherms for 135 soil
horizons from 36 profiles of the major forest soils of the temperate zones
(Leptosols, Vertisols, Cambisols, Luvisols, Podzols, Stagnosols, and Gley-
sols). When solutions containing no DOMwere added, the release of DOM
was greatest for horizons rich in organic C. In subsoil horizons,
DOM release was much less. Most of the topsoil horizons showed weak
DOMsorption. This was caused by poor concentrations of sorbents (clay and
sesquioxides) and/or high concentrations of organic C. Organic C appar-
ently decreased DOM sorption by occupying binding sites. Subsoils rich in
clay and sesquioxides showed a strong retention of DOM. The majority of
the soils preferentially sorbed hydrophobic DOM—caused by the greater
affinity of hydrophobic DOM to oxide/hydroxide soil constituents. From
microcalorimetric, FTIR, and 13C NMR analyses, Gu et al. (1994) con-
cluded that ligand exchange between carboxyl/hydroxyl formational
groups of the SOM and iron oxide surfaces were the dominant sorption
mechanisms, especially under acidic or slightly acidic pH conditions.
26 Nanthi S. Bolan et al.
In deeper mineral soil horizons of forest lands, DOM fluxes declined
from 10–40 g C m�2 yr�1 translocated from the organic surface layer into
the mineral soil horizons to about 1–10 g m�2 yr�1 in deeper mineral
horizons, indicating substantial retentions of DOM in subsoil horizons
(Guggenberger and Kaiser, 2003). This observation and other similar obser-
vations prompted a general hypothesis that retention of DOM in the soil (or
sediment) solid phase is a mechanism that promotes stability and conserva-
tion SOM in soils (Hedges and Oades, 1997; Kaiser and Guggenberger,
2000; Kaiser et al., 1996; Michalzik and Matzner, 1999; Sollins et al., 1996).
However, Guggenberger and Kaiser (2003) estimated a mean subsidence
time of sorbed SOM of about 4–30 years, inferring that instead of the
“preservation” role of sorbed DOM, such DOMmay enhance bioavailabil-
ity to microbe causing subsequent biodegradation.
Investigative consensus indicates that high organic C concentrations of
the soil decrease DOM sorption, especially the hydrophilic fraction. In soils
free of carbonates, sorption is related to oxalate-extractable Al and dithio-
nate-extractable Fe; however, in carbonitic soils, DOM sorption is corre-
lated with dithionate-extractable Fe only. The sorption of DOM by topsoil
is always less than in subsoil samples. Sorption is generally high in B horizons
of Alfisols, Inceptisols, and Spodosols with low organic C content and high
contents of oxalate-extractable Fe and Al and dithionate-extractable Fe,
whereas little or no sorption is noticed in soils with high contents of organic
C and/or low contents of metal oxides alone, much as those in the A and E
horizons ( Jin et al., 2008; Kaiser et al., 1996; Kothawala et al., 2009; Muller
et al., 2009)
Dissolved organic C concentrations in soil solutions can be as low as
0.1–3.6 mmol dm�3 in forest soils often in contact with subsurface horizons
(Cronan and Aiken, 1985; Guggenberger and Zech, 1993; Laik et al., 2009;
Laudon et al., 2009; McDowell and Likens, 1988; Sanderman et al., 2008).
The decrease in DOM concentrations is characterized by a change inDOM
composition, indicated by a preferential decrease of the hydrophobic frac-
tion (Guggenberger and Zech, 1993). This was subsequently confirmed by
Kaiser et al. (1996) where the majority of the soils studied preferentially
sorbed hydrophobic DOM, apparently caused by higher affinity of the
hydrophobic fraction for metal oxides/hydroxides in the soil matrix.
Indeed, sorption of hydrophobic DOM by some soils was accompanied
by the release of hydrophilic substances (Moore andMatos, 1999; Ussiri and
Johnson, 2004).
The formation of soil organo-mineral complexes is a key reaction in the
carbon cycle in soil, since organic materials acquire a resistance to decom-
position due to the formation of the complexes. Adsorption of DOM onto
soil minerals provides a model of this important process. Adsorption of
DOM onto samples from Andisols, Inceptisols, and Entisols in batch
experiments was compared in terms of the quantitative relationship
Dissolved Organic Matter 27
between the soil properties and the adsorption behavior of DOM (Nambu
and Yonebayashi, 2000). Andisols showed a particularly high efficiency of
adsorption compared with those from other soils that contained a compara-
ble amount of organic carbon. Although the adsorption mechanisms varied
among soils, two soil variables, the degree of carbon accumulation in the soil
sample (or total carbon/specific surface area ratio), and the amount of ligand
exchange sites on labile aluminum accounted for the variation in DOM
adsorption.
In general, DOM components that are low in molecular weight, organic
N, acidic groups, and aromatic structures can be expected to remain soluble
in the soil’s aqueous phase, whereas constituents that are rich in organic N,
acidic groups, and with high aromaticity can be preferentially sorbed (Gu
et al., 1995; McKnight et al., 1992). While the sorption of DOM in soils
increases with increasing levels of Fe and Al oxides in soils, it decreases with
increasing concentrations of organic matter.
4.2. Biodegradation
Biological assimilation of organic carbon and subsequent generation of
DOM plays an important role in controlling DOM dynamics in soils
(Figure 1). Thus, DOM originates primarily from the decomposition of
SOM that had accumulated through vegetation, the addition of biological
waste materials (e.g., biosolids and livestock manures), the release of root
exudates, and the lysis of microorganisms. The decomposer community in
soil consists of a wide range of bacteria, fungi, protista, and invertebrates
(Dilly et al., 2004; Kalbitz et al., 2000; Swift et al., 1979). Considerable
emphasis has focused on microorganisms because of their dual roles as
decomposition agent and as a sink of labile organic C. Microbial assimilates
represent an important source of DOM released from the forest floor, while
microbial biomass serves as an important reservoir of DOM. Soil fauna,
including earthworms, can facilitate the turnover rate of microbial biomass
in soil (Aira et al., 2008; Kalbitz et al., 2000; Osler and Sommerkorn, 2007;
Siira-Pietikainen and Haimi, 2009).
Dissolved organic carbon is an important substrate for microorganisms
(Marschner and Bredow, 2002; Michelsen et al., 2004; Qualls, 2005).
Laboratory incubation studies of varying length have indicated that
10–44% of DOM in soil solution is microbiologically degradable ( Jandl
and Sletten, 1999; Kalbitz et al., 2000; Qualls, 2005; Sachse et al., 2001).
The more labile fraction of DOM is more readily mineralized or assimilated
into microbial biomass (Nelsen et al., 1994; Qualls, 2005).
It is likely that DOM production is controlled by the same factors
controlling biological activity. The decomposition rate of DOM is influenced
by soil depth, land use, soil fertility, etc. It decreased with increasing soil depth
and is less in forest than in arable soils (Ludwig et al., 2000;Muller et al., 2009).
28 Nanthi S. Bolan et al.
Simply,microbial activitywith depth is limited by the bioavailability of organic
C as a substrate (Celerier et al., 2009; Ghiorse and Wilson, 1988; Rodriguez-
Zaraqoza et al., 2008; Zablotowicz et al., 2009) or the supply of essential
nutrients such asN and P. It is well known thatCdecomposition rate decreases
with decreasing available N (Chantigny et al., 1999; Enowashu et al., 2009;
Frank and Groffman, 2009; Sirulnik et al., 2007).
Most of the DOM in soils is the end product of microbial metabolism of
organic residues. Fresh litter also contributes significantly to the production
of DOM, indicating the presence of DOM in the original litter. Ludwig
et al. (2000) studied the production of DOM in soils from two sites with
different microbial activities using C13-depleted plants of differing decom-
posability (Epilobium angustifolium and Calamagrostis epigeios). Cumulative
DOM production was markedly greater in the readily decomposing Epilo-
bium experiment (2% of the added C) than in the slow decomposing
Calamagrostic experiments (0.1% of the added C).
The rate of biodegradation of DOM varies among sources, which has
been attributed to the difference in its composition (Kalbitz et al., 2000).
Some of the hydrophobic compounds extracted in the DOM are less
accessible to microbial degradation than hydrophilic compounds (Amon
et al., 2001; Kalbitz et al., 2003; Qualls, 2005). Based on biodegradation
kinetics, DOM in soils is grouped into various categories such as labile and
nonlabile fractions (Marschner and Kalbitz, 2003; Saadi et al., 2006).
Microbes selectively degrade the less recalcitrant compounds, thereby
gradually increasing the average recalcitrance of the remaining organic
carbon (Bowen et al., 2009; Sollins et al., 1996; Waldrop and Firestone,
2004). 14C-dating has indicated that organic C in deeper horizons had
longer residence times, indicating lower bioavailability to microbes (Chiti
et al., 2009; Favilli et al., 2008; Oades, 1984).
In summary, while microbial degradation of SOM, followed by desorp-
tion of organic substances from the soil matrix and leaching of soluble
organic compounds from fresh litter are viewed as the most important
processes causing the release of DOM (Currie et al., 1996; Guggenberger
et al., 1994a; Marschner and Kalbitz, 2003; Qualls and Haines, 1991),
microbial assimilation of readily available carbon in DOM results in the
ultimate degradation of DOM to carbon dioxide.
4.3. Photodegradation
Although DOM undergoes photochemical and microbial degradation, the
former process dominates in aquatic systems and the latter on land
(Marschner and Kalbitz, 2003; Minor et al., 2007; Mostofa et al., 2007).
In Fe-rich surface waters, light-induced redox cycling of Fe and DOM
photo-oxidation are strongly coupled (Norton et al., 2008; Shiller et al.,
2006). Iron can catalyze DOM photo-oxidation via ligand-to-metal(loid)
Dissolved Organic Matter 29
charge transfer reactions of Fe(III)–DOM complexes and through DOM
oxidation by the hydroxyl radical (HOo) formed via a Fenton reaction
(Giroto et al., 2006; Voelker et al., 1997).
Photo-oxidation can enhance the turnover of DOM in aquatic systems,
transforming labile into more recalcitrant (less bioavailable) components and
vice versa (Benner and Biddanda, 1998; Obernosterer et al., 1999). For
example, photo-cleavage and photo-oxidation of HMW DOM resulted in
the release of bioavailable LMW that stimulated bacterioplankton activity
(Keiber et al., 1989; Mostofa et al., 2007; Wetzel, et al., 1995).
Pullin et al. (2004) observed that photo-irradiation of DOM decreased
the abundance of HMW components and formed new LMW components,
including LMW carboxylic acids, that is, acetic, formic, and malonic acids.
This can alter the complexation potential of DOMwith metal(loid)s such as
Fe. For example, it has been shown that intermediate and/or HMW, more
aromatic, components of DOM sorb preferentially onto Fe(III) oxyhydr-
oxide surfaces (Gu et al., 1995; Kaiser and Zech, 1997; Kothawala et al.,
2009; Meier et al., 1999; Vandenbruwane et al., 2007; Zhou et al., 2001).
Thus, by decreasing