The Science of the Total Environment 293 (2002) 1–29
The need for comprehensive and consistent treatment of the
nitrogen cycle in nitrogen cycling and mass balance studies: I.
Terrestrial nitrogen cycle
E.C. Krug*, D. Winstanley
Illinois State Water Survey, 2204 Griffith Drive, Champaign, IL 61820, USA
Received 1 October 2001; accepted 1 December 2001
Abstract
A review of conceptual models that scientists use to characterize the nitrogen (N) cycle and to conduct N mass
balance studies at global, regional and local scales is presented. Large uncertainties in processes and process rates
make it difficult to conduct precise N mass balances and the dominant conceptual model has changed in recent
decades. An earlier conceptual model recognized explicitly that human activities, especially agriculture, have both
depleted terrestrial N and increased the fixation of atmospheric N in biologically available forms. The current
conceptual model does not include adequate treatment of the depletion of the terrestrial N reservoir, the resulting
transfer of N to the hydrosphere and atmosphere, or the cycling of terrestrial N below the plow layer. Thus, it delivers
an unrealistically limited view of human influences on the N cycle. It is recommended that a comprehensive and
consistent treatment of terrestrial N cycling be developed to better facilitate scientific explanation of historical Nrelated environmental changes and more closely balance N budgets on a range of geographical and temporal scales.
Improved N-cycle models will provide an improved scientific basis for answering important resource management
and policy questions. 䊚 2002 Elsevier Science B.V. All rights reserved.
Keywords: Nitrogen; Biogeochemical cycles; Mass balance; Mississippi River Basin; Illinois
1. Introduction
Conceptual models provide a basic mode for
organizing scientific data and human thought and
are necessary for the development of mathematical
models. When models are widely accepted, they
can shape scientific research agendas, influence
*Corresponding author. Tel.: q1-217-244-0877; fax: q1217-333-4983.
E-mail address: [email protected] (E.C. Krug).
the public’s view of the world and provide a
framework for answering important policy and
resource management questions. The effectiveness
of these policies and strategies is, in part, dependent upon the validity and robustness of data and
models.
Nitrogen (N) is an element essential for life. In
many biological communities growth is limited by
insufficient biologically available N. Nevertheless,
in recent decades anthropogenic addition of N has
0048-9697/02/$ - see front matter 䊚 2002 Elsevier Science B.V. All rights reserved.
PII: S 0 0 4 8 - 9 6 9 7 Ž 0 1 . 0 1 1 3 3 - 0
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E.C. Krug, D. Winstanley / The Science of the Total Environment 293 (2002) 1–29
come to be often regarded as undesirable. Objections range from the purely philosophical to the
objectively tangible: from the ‘unbalancing’ of the
implicit biological and chemical perfection of the
undisturbed natural state (Kinzig and Socolow,
1994; Tilman and Downing, 1994; Jordan and
Weller, 1996; Vitousek et al., 1997; Matson et al.,
1999; Reich et al., 2001) to increasing the amounts
of N contributing to smog episodes, global warming, stratospheric ozone depletion and acid rain in
the atmosphere, eutrophication, hypoxia, water-use
impairment and violations of drinking-water standards in the hydrosphere (National Research Council, 2000). Since the philosophy of worldview
informs that of the scientific and the technical
communities, even these ostensibly objective arenas are conditionally defined (e.g. Kuhn, 1970).
The stated goal of the United States of America’s
(USA’s) 1972 Clean Water Act wP. L. 92-500, Sec.
101. (a)x is to achieve zero discharge of pollutants,
including N, to the waters and surrounding oceans
of the contiguous states. The standard of the
undisturbed natural state is a reference state the
United States Environmental Protection Agency
(USEPA) is using to provide technical guidance
for the development of N criteria for water bodies.
In Illinois, estimated presettlement biological conditions were used by the Critical Trends Assessment Program (CTAP, 2001) as the yardstick to
assess ecosystem health. Hence, the philosophy of
worldview makes it scientifically and technically
important to establish measures of the undisturbed
natural state and to use these measures to define
and quantify the impacts of human activities on
the environment.
In addition to recognizing the influence that
underlying values have in defining environmental
problems, we need to recognize the influence that
the various chemical forms of N, transformation
rates, pathways and residence times of N in the
environment have in impacting the environment.
Dissolved and particulate forms of inorganic
(ammonia, ammonium, nitrite and nitrate) plus
organic N in water are defined as total N. It is
widely recognized on the bases of science and
value judgment that concentrations and fluxes of
total N need to be reduced in order to improve
water quality and the health of aquatic ecosystems
(CENR, 2000; USEPA, 2000). In this regard, two
examples of long-standing, N-related policy questions in the USA can be given. Is the N content
and fertility of agricultural soils still being depleted? How far can the concentration and fluxes of
N in Midwest streams be reduced while enhancing
the viability of agriculture? The answers to these
questions depend, in part, on sound scientific
understanding of N cycling and the strength of
causal source–receptor chains.
Consistent with the generic concept of biogeochemical cycling, scientific understanding and
descriptions of N dynamics are embedded in the
concept of N cycling. Inherent in the concept of
N cycling is the principle of the conservation of
mass. This principle allows a mass balance to be
conducted so that a mass of N can, at least in
principle, be tracked and accounted for as it is
transformed into a variety of chemical forms during N cycling within and among N reservoirs. A
pre-requisite for balancing the inputs and outputs
of N in a watershed, for example, is to identify all
significant sources and sinks of N, all N cycled
within N reservoirs, and all N fluxes among the
reservoirs. In reality, the N cycle is extremely
complex, and it is difficult to conduct a precise N
mass balance on any geographical scale. Typically,
uncertainties and gaps in knowledge are large.
Measuredymodeled outputs do not equal inputs on
any scale and balance has to be achieved by fitting
(e.g. Allison, 1955; Rosswall, 1976; LaRue and
Patterson, 1981; Stevenson, 1959a; David et al.,
1997; Goolsby et al., 1999; Jenkinson, 2001),
examples of which will be later given.
The focus of this study is the terrestrial N
reservoir: N inputs to this reservoir, N content and
cycling within this reservoir and the flux of N
from this reservoir to the hydrosphere. From a
literature review, terrestrial N cycling appears to
be one of the greatest sources of uncertainty in Ncycle and N-mass-balance studies. An additional
study of the N cycle and N mass balance in the
hydrosphere has been made. It will be published
as a follow-up article to broaden analysis of the
soundness of understanding of N cycling and to
test the strength of causal source-receptor chains
to the endpoint of watershed N export.
E.C. Krug, D. Winstanley / The Science of the Total Environment 293 (2002) 1–29
3
Fig. 1. Geographical scales and locations of the studies.
2. Methods
The approach used in this study was to identify
and evaluate (a) historical N data, and (b) the
evolution of conceptual N-cycle models at global,
regional and local scales. Descriptions over the
past 200 years at the global scale provide a broad
framework for more detailed N-cycling and Nmass-balance studies at regional and local scales
and on shorter time scales. The global 200-year
scale also provides a framework for understanding
the natural N cycle and human influences on it
(Galloway et al., 1995). The Mississippi-Atchafalaya River Basin (MARB), which drains more
than 40% of the contiguous continental USA into
the Gulf of Mexico (Fig. 1), was selected to
illustrate conceptual N cycling and N-mass-balance
modeling at the regional scale. The MARB is a
breadbasket of the world. It contains 76%, 101.5
million hectares (ha), of the USA’s 132.8 million
ha of cropped land (Doering et al., 1999, Table
4.1–3). Increased use of chemical-N fertilizer in
the MARB since the 1950s are reported to be
principally responsible for flooding the MARB
with N — saturating the N cycle and N reservoir
of the drainage basin. The N in excess of the Nsaturation capacity of the MARB has been draining
to the Northern Gulf of Mexico and has increased
algae production enough to create an oxygendepleted ‘dead zone’ (CENR, 2000). Illinois, a
major agricultural state in the MARB (Fig. 1),
was selected to illustrate conceptual modeling at
the state level. And the headwaters of the Embarras
River Watershed (ERW), an agricultural watershed
in central Illinois, were selected to illustrate conceptual modeling at the local scale (Fig. 1).
In reviewing data and literature, the authors
determined that there now appears to be no comprehensive and consistent conceptual model of
terrestrial N cycling in N models and N-massbalance studies, or in evaluating the impacts of
human activities on the N cycle. In order to provide
a framework for documenting the lack of a consistent treatment of terrestrial N cycling and for
making recommendations for improvements, the
authors identified two quite different conceptual
models of the N cycle. Because the two models
appear to have developed at different periods of
time, they are called the current conceptual model
(current model) and the previous conceptual model
(previous model). The current model appears to
have been developed mainly over the last three
decades, while the previous model developed
mainly prior to this time. Whereas a change of
scientific perspective and understanding over time
is often a progressive step, this study questions
whether the current model offers an improvement
over the previous model. There is some overlap in
time and content between the two models, but this
report focuses on (a) documenting the main characteristics of the models and related data, and (b)
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E.C. Krug, D. Winstanley / The Science of the Total Environment 293 (2002) 1–29
identifying the major differences between the models. This analysis lays the foundation for recommending improvements in the treatment of
terrestrial N cycling and mass-balance studies. A
subsequent article will do likewise for analysis of
the aquatic N cycle.
3. Terrestrial nitrogen
Of the earth’s estimated 1.6=1017 metric tons
of N, the geosphere contains 98%; the atmosphere,
2%; and the biosphere, 0.0002% (Delwiche, 1970;
Stevenson, 1982b, p. 4). Under the current conceptual model, a focus of environmental concern
is on the effects that an estimated 150 million
metric tons of anthropogenic ammonium (NH4-N)
fixation yry1 may have on the biosphere (Kinzig
and Socolow, 1994; Galloway et al., 1995; Vitousek et al., 1997; Galloway, 1998). Under a previous
conceptual model, such concern had focused on
both anthropogenic N additions to and subtractions
from the biosphere. It was well recognized that
human activities had decreased the size and magnitude of cycling of the earth’s terrestrial (soil) N
reservoir (King and Whitson, 1901, 1902; Conner,
1922; Albrecht, 1938; Jenny, 1941; Viets and
Hageman, 1971; Rosswall, 1976). Indeed, concern
about agronomically induced losses of soil N was
a major driving factor behind the conservation
movement (Lawes and Gilbert, 1863; Storer, 1905;
Hopkins, 1910; Russell and Richards, 1920; Conner, 1922; Albrecht, 1938; Bennett, 1939; Jenkinson, 1991); e.g. ‘The conclusion that virtually all
the organic matter (and N) might eventually be
lost from soil by cropping caused considerable
alarm in agronomic circles. It was feared that
unless drastic measures were taken to maintain
organic matter reserves, many soils would become
unproductive.... For soils of the wCxorn wBxelt,
approximately 25% of the N was found to be lost
the first 20 years, 10% the second 20 years and
7% the third 20 years’ (Stevenson, 1986, p. 55).
The historic loss of carbon (C) associated with
soil organic matter (SOM) loss due to anthropogenic activities is recognized and accounted for in
the current conceptual model of the global C cycle.
However, the historic loss of N associated with
SOM loss due to anthropogenic activities is not
well recognized and accounted for in the current
conceptual N-cycle model at any geographical
scale (Aber et al., 1989; Kinzig and Socolow,
1994; Galloway et al., 1995; Jordan and Weller,
1996; David et al., 1997; Goolsby et al., 1999;
CENR, 2000; National Research Council, 2000).
Historic soil N depletion is potentially significant
relative to the direct effect of anthropogenic NH4 –
N fixation because of the relatively great size of
the soil N reservoir, the rates at which decomposition processes mineralize this organic N reservoir
to produce NH4 –N for plant and microbial uptake,
microbial nitrification and denitrification, and the
rates at which various forms of soil N may be
lost, especially to the hydrosphere. The size of the
global soil N reservoir is estimated to be ;240 000
million metric tons of N just for the upper 1-meter
(m) of soil. Approximately 90% of this soil N
reservoir is estimated to be organic N (Stevenson,
1982b, pp. 4–6). Most commonly, the rate of
production of inorganic N by mineralization of
soil organic N is reported as net mineralization —
the difference between the total amount of organic
N mineralized to inorganic N (gross N mineralization) minus that amount of inorganic N immobilized, i.e. that mineralized N which is rapidly
utilized and thereby, converted back to organic N
(Jansson and Persson, 1982). Thus, gross N mineralization is always larger than net N mineralization (Jansson and Persson, 1982; Smith et al.,
1994; Takahashi, 2001), and studies have shown
that rates of gross N mineralization in soils are
approximately twice those of net N mineralization
(e.g. Paul and Juma, 1981; Schimel, 1986; Wang
et al., 2001).
Rosswall (1976) estimated that globally mineralization of this soil organic N to NH4 –N, produced 5800 million metric tons NH4 –N yry1
available for plant and microbial uptake, nitrification, denitrification and leaching. Soil N mineralization values are less now than they would have
been 60–70 years earlier because agricultural
activities were estimated to have reduced soil N
by 13 000 million metric tons. This loss would
have preferentially been of the most rapidly mineralized soil N (Rosswall, 1976; Krug and Winstanley, 2000). Using a 1–3% net N mineralization
range of rates currently used for agricultural soils
E.C. Krug, D. Winstanley / The Science of the Total Environment 293 (2002) 1–29
(David et al., 1996; Goolsby et al., 1999, p. 48),
and assuming no change in net mineralization rate,
agricultural activities have resulted in a global
decrease of 150–450 million metric tons NH4-N
yry1 in the top 1 m of soil, with a best estimate
of ;300 million metric tons NH4-N yry1 net
decrease using the best estimate of net mineralization of Goolsby et al., (1999, p. 48), an ;600
million metric tons NH4 –N yry1 gross N mineralization decrease. Regarding the effects of agriculture on the soil N reservoir in the USA,
agricultural activities were estimated to have
reduced the soil N reservoir of the contiguous
continental USA by 1.75 billion tons (Viets and
Hageman, 1971, p. 14). Repeating the above
analysis, agricultural activities have resulted in a
net decrease of ;35 million metric tons NH4 –N
yry1, ;70 million metric tons NH4 –N yry1 gross
N mineralization decrease for the soils of the
contiguous continental USA. This value can be
compared to a high of approximately 11 million
tons N yry1 of chemical-N fertilizer added to USA
agricultural soils in recent years (Association of
American Plant Food Officials, 2000, Table 1).
Since before the post-World War II era of
chemical-N fertilizer it was recognized that, on
average, approximately 20–30% of the mineralized
N lost from cropped soils was lost to the hydrosphere as NO3 –N (e.g. Storer, 1905, volume 3, p.
341; Viets and Hageman, 1971; Rosswall, 1976;
Stevenson, 1986, pp. 145–149). Depletion of soil
N also resulted from volatilization losses to the
atmosphere, leaching of soluble organic N from
soil organic matter, animal wastes and plant materials and the transport of particulate organic N
(erosion) to the hydrosphere in the forms of soil
and plant debris (Riddell, 1846; Kofoid, 1903;
Palmer, 1903; Hopkins, 1910; Conner, 1922;
Albrecht, 1938; Watkins, 1943; Allison, 1955;
Koelling and Kucera, 1965; Tukey, 1966; Timmons
et al., 1968; White, 1973a,b; Rosswall, 1976;
Timmons and Holt, 1977; Stevenson, 1982a; Thurman, 1986, pp. 76–80, 151–180; Northrup et al.,
1995; Elrashidi et al., 1999). Indeed, review of the
literature indicates that erosion of soil organic N
was the most important source of N delivery to
surface waters from agricultural lands in the USA.
And, although rates of erosion have decreased
5
greatly in recent decades, due largely to improved
conservation practices (e.g. Stevenson, 1986, p.
148), erosion is still reported as being the most
important source of N loss from cropped land to
waters in the MARB (Doering et al., 1999, pp.
27–30, Table 4.1–1). Once in the hydrosphere,
inorganic and organic forms of terrestrial N can
undergo interconversion and further cycling and
recycling (Kofoid, 1903; Palmer 1903; Fair et al.,
1941; Reid, 1961, pp. 183–187; Feth, 1966; Jewell, 1971; Miller et al., 1974; Ruttner, 1974, pp.
87–91; Wetzel, 1975, pp. 186, 214, 564–565;
Stewart et al., 1982; Harrison, 1983; Timperley et
al., 1985) as can the large amounts of N fixed
within the hydrosphere itself (e.g. Allison et al.,
1937; Dugdale and Neess, 1961; National
Research Council, 1969; Stewart, 1969; Brouzes
et al., 1969; Granhall and Lundgren, 1971; Ogan,
1979; Wetzel, 1983, p. 567; DeLaune et al., 1986;
Paerl, 1990; Chu and Alvarez-Cohen, 1998; Currin
and Paerl, 1998).
The importance of quantifying all sources of
organic N can be demonstrated in the MARB. The
USA’s National Centers for Coastal Ocean Science
Gulf of Mexico Hypoxia Assessment (Hypoxia
Assessment) reported that 37% of the mean annual
flux of N from the MARB to the Gulf of Mexico
in 1980–1996 was directly as organic N (Goolsby
et al., 1999, p. 13). In a mass-balance discussion
(pp. 63–64), the authors stated that SOM contributed only to inorganic N inputs and outputs
through the process of mineralization. Soil organic
N was not identified as a direct input. Mineralized
soil organic N was estimated to contribute approximately 30% of the total N inputs to the MARB
annually since 1980, similar to the current annual
input from fertilizer and to contribute significantly
to N export from the MARB (p. 66). The large
flux of organic N was not explained.
4. Results
4.1. The current conceptual model
The current model defines the earth’s unreactive
N2 gas reservoir (the atmosphere) as the source of
reactive biospheric N and the cycling of biospheric
N is regarded as largely synonymous with total N
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E.C. Krug, D. Winstanley / The Science of the Total Environment 293 (2002) 1–29
cycling. The magnitude of the global N cycle is
defined as the amount of N2 gas fixed annually
and the amount of N gases returned annually to
the atmosphere, principally by microbial nitrificationydenitrification in the biosphere, geosphere and
hydrosphere.
The consensus is that prior to the Industrial Era
an estimated ;130 million metric tons N2 yry1
were fixed from the atmosphere on the continents,
primarily by microorganisms, and ;130 million
metric tons of this fixed N were returned annually
to the atmosphere as gases, principally by microbial nitrificationydenitrification (Kinzig and Socolow, 1994; Galloway et al., 1995; Smil, 1999
National Research Council, 2000). Studies of N
cycling at the global scale typically include uncertainties in process rates of up to an order of
magnitude (Soderlund and Svensson, 1976; Galloway et al., 1995; Vitousek et al., 1997; Galloway,
1998; Jenkinson, 2001). Recognizing our poor
understanding of the many N fluxes and their
considerable variations, Smil (1999) appropriately
presented inputs and losses as ranges rather than
as single values.
The current model recognizes that large amounts
of biologically useful N are stored and cycled in
the terrestrial N reservoir, but internal cycling of
N within the biosphere, geosphere and hydrosphere
is included only to the extent of annual cycling of
inputs. Under the current model it is asserted that
human activities, by increasing the fixation of
atmospheric N into biologically available forms
(primarily through fertilizer production, biological
fixation and fossil fuel combustion since ;1950),
have approximately doubled the magnitude of the
global biospheric N cycle on the continents to an
estimated ;280 million metric tons yry1 (Kinzig
and Socolow, 1994; Galloway et al., 1995; Vitousek et al., 1997; Smil, 1999; National Research
Council, 2000, p. 97); e.g. ‘the supply of reactive
N to global terrestrial ecosystems has doubled as
a consequence of human activity’ (Galloway,
1998, p. 15).
The foundation of the current model is the
implicit biological and chemical perfection of the
undisturbed natural state. Under the current model,
in the absence of human activity, the terrestrial N
cycle is balanced and efficient, losing but little N
to the hydrosphere (Howarth et al., 1996; David
et al., 1997; Downing et al., 1999, p. 13; Goolsby
et al., 1999, p. 31; National Research Council,
2000, pp. 100–101); e.g. ‘In a balanced natural
system the amounts of organic nitrogen and nitrate
dissolved in the water remain low, the population
of algae and animals is correspondingly small, and
the water is clear and pure. And because the
natural nitrogen cycle in soil is tightly contained,
relatively little nitrogen is added to the water in
rainfall, or in drainage from the land’ (Commoner,
1970, p. 73). In the absence of human activities,
reactive N was retained by terrestrial ecosystems
and hydrologic losses from regions of N fixation
were minimized. With the additional anthropogenic N inputs, biospheric systems are modeled as
becoming N saturated — receiving more N than
they can handle. Like the run-off from a heavy
rain, the soils of N-polluted terrestrial ecosystems,
unable to handle the overload of N, leak much of
the excess N to the hydrosphere (Aber et al., 1989;
Kinzig and Socolow, 1994; Galloway et al., 1995;
Howarth et al., 1996; Jordan and Weller, 1996;
David et al., 1997; Downing et al., 1999, p. 13;
Goolsby et al., 1999, p. 31; Smil, 1999; National
Research Council, 2000, pp. 100–101).
Prior to human disturbance approximately 10 to
30% of the reactive N formed by terrestrial N
fixation was estimated to have been lost to coastal
oceans via the rivers (Galloway et al., 1995, p.
236). With the doubling of terrestrial reactive N
inputs by combined natural and anthropogenic N
fixation, the release of reactive N from the continents to coastal oceans via rivers was reported to
have doubled (Galloway et al., 1995, p. 243). That
the reported percent of leakage to the hydrosphere
has remained essentially the same means that the
terrestrial ecosystem is still processing inputs of
reactive N with the same efficiency as it did before
there were any anthropogenic inputs of reactive N.
This is internally inconsistent with the conclusion
that the terrestrial ecosystem has become N saturated, cannot process the N in excess of its saturation point and is, therefore, leaking this excess
N.
Under the current model, human activities are
considered as increasing the N content of the
terrestrial N reservoir, which eventually forces the
E.C. Krug, D. Winstanley / The Science of the Total Environment 293 (2002) 1–29
conclusion of N saturation of the terrestrial environment — the accumulation of reactive N in the
terrestrial environment beyond what the terrestrial
environment can assimilate. It is now reported that,
‘The single largest global change in the nitrogen
cycle comes from increased reliance on w83 million
metric tons yry1x synthetic inorganic fertilizers...’
(National Research Council, 2000, p. 97). However, billions of metric tons of inorganic N are
naturally bioavailable in soils each year (Viets and
Hageman, 1971; Rosswall, 1976; Stevenson, 1986,
pp. 155–215). Thus, while 150 million metric tons
of reactive N yry1 are asserted to represent approximately a doubling of the supply of reactive N to
the terrestrial biosphere (Kinzig and Socolow,
1994; Galloway et al., 1995; Vitousek et al., 1997;
Galloway, 1998; Smil, 1999; National Research
Council, 2000, p. 97), such an input falls far short
of a doubling of the supply. Furthermore, as
previously discussed, anthropogenic activities,
principally agriculture, have reduced the amount
of inorganic N annually made bioavailable by a
net of ;300 million metric tons. Scientists operating under the current model do not acknowledge
that anthropogenic activities, because they both
add to and subtract N from the environment, have
had the apparent net effect of decreasing the
amount of reactive N available to the global
terrestrial biosphere.
On a regional scale (Fig. 1), the Integrated
Assessment of Hypoxia in the Northern Gulf of
Mexico (CENR, 2000) reported that the concentration of N in rivers in the MARB and N fluxes
from these rivers to the Gulf of Mexico have
increased significantly in the last 100 years. The
most significant trends in N inputs and outputs are
reported to have occurred between the 1950s and
1980s, when N fertilizer input more than tripled.
The CENR (2000) reported N inputs and outputs for the MARB for the period 1951 to 1996.
The estimated average annual N output from the
MARB was approximately equal to the annual N
input for the 1980–1996 period — on the order
of 21 million metric tons N yry1 (p. 66). Although
the Hypoxia Assessment (Goolsby et al., 1999;
Brezonik et al., 1999, pp. 27–28) and the CENR
(2000) recognized a close positive correlation
between precipitation and NO3 –N production and
7
fluxes, and recognized that precipitation had
increased in the MARB since the 1950s, they did
not quantify the extent to which increased NO3 –
N production and fluxes from the MARB were
due to increased precipitation, rather than increased
N inputs (Goolsby et al., 1999, p. 71). The
Hypoxia Assessment stated that hypoxia in the
northern Gulf of Mexico has intensified only since
the 1950s, and that increased use of N fertilizer is
the major cause (Goolsby et al., 1999; CENR,
2000), even though it is also reported that total N
and NO3 –N from net soil-N mineralization and
from chemical-N fertilizers are indistinguishable
(Goolsby et al., 1999, pp. 67–73). The Hypoxia
Assessment reported that legumes, e.g. soybean,
generally are not net contributors of N to the soil
system (Goolsby et al., 1999, p. 62), and N
removal in harvested crops and pasture was identified as being nearly 50% larger than fertilizer
inputs (p. 66).
Neither the Hypoxia Assessment (Goolsby et
al., 1999) nor the CENR (2000) provided an
analysis of N inputs and outputs for the MARB
prior to the 1950s. The Hypoxia Assessment did
recognize that the total N content of agricultural
soils averages ;333 000 kilograms (kg) N per
kilometer2 (kmy2) in the upper 30 centimeters
(cm) of soil and that net mineralization can be as
high as 40 000 kg N kmy2 yry1 in virgin cultivated
land (Goolsby et al., 1999, pp. 47–48). Calculation based on these figures yields a net mineralization rate of soil organic N of approximately 12%
yry1, given today’s soil-N content. The Hypoxia
Assessment also recognized that soil fertility was
reduced by the mining of soil nutrients (Goolsby
et al., 1999, p. 44), and that some of the N in soil
organic matter is lost through erosion (p. 52).
However, the Hypoxia Assessment did not directly
quantify and incorporate any of these N losses in
its MARB N budget (Goolsby et al., 1999).
Neither did the Hypoxia Assessment (Goolsby et
al., 1999) or the CENR (2000) discuss the implications of these N losses for depletion of the soilN reservoir, N leaching rates, N concentrations
and fluxes in the rivers, or N mass balances.
Analysis of the MARB indicates that the operations conducted under the current model confuse
the N cycle on the regional scale in such a way
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E.C. Krug, D. Winstanley / The Science of the Total Environment 293 (2002) 1–29
as to make the research findings consistent with
the N-saturation hypothesis. The Hypoxia Assessment noted that net ‘Mineralization of the soil
organic N is the largest source (;75%) of recycled
N, or N that was already in the system’ (Goolsby
et al., 1999, p. 66). Having said this, the Hypoxia
Assessment and the CENR went on to list net soilN mineralization as an external input into the
MARB system. In its input-output budgets and N
mass balance studies, the Hypoxia Assessment and
the CENR treated net soil-N mineralization as an
external input to the MARB by which the terrestrial N reservoir of the MARB, as shown by the
cumulative residuals, was calculated to have grown
by nearly 60 million metric tons of N since 1955
(Goolsby et al., 1999, Fig. 6.4B, p. 65; CENR
2000, p. 31). On the other hand, by being consistent with the fact that net mineralization of soil N
is net mineralization of N already in the system,
Hypoxia Assessment data (Goolsby et al., 1999,
Fig. 6.4A, pp. 64–65, Table 5.2) indicate that the
MARB has lost ;220 million metric tons of N
since 1955. And, as stated above, this 220 million
metric ton loss of N is in addition to the large
quantities of N lost from the MARB earlier in the
19 and 20th centuries and from the ongoing
processes of erosion and sediment transport.
Goodness of fit between inputs and outputs and
the achievement of mass balance for the MARB
since 1955 appear to validate the current model
and its fundamental premise of N saturation. However, the goodness of fit between MARB N inputs
and outputs, and the N mass balance achieved by
growth of the MARB’s terrestrial N reservoir since
1955 are fortuitous. In addition to net mineralized
N being misclassified as an external input to the
MARB, the magnitude of net mineralized N was
miscalculated is such a way as to underestimate
net soil N mineralization. Namely, by defining net
mineralization of N in SOM from the mass of
SOM in the top 30 cm (12 inches) with a SOMN content of 3%, a net mineralization rate of 2%
yry1, and only for MARB cropped land (Goolsby
et al., 1999, p. 48) the amount of net N mineralization of N in SOM of the MARB was greatly
underestimated.
First, the estimate of net N mineralization in the
MARB is derived from cropped land only. Whereas
other N inputs are considered for the whole of the
;300 million ha MARB (Goolsby et al., 1999;
Mitsch et al., 1999, p. 1), net soil-N mineralization
is underestimated in the N budget by assuming
that net mineralization of soil N only occurs in the
;100 million ha of cropped land within the ;300
million ha MARB (Goolsby et al., 1999, pp. 48,
64; Doering et al., 1999, Table 4.1–3; Mitsch et
al., 1999, p. 1).
Second, underestimation of net soil mineralization is suggested by Hypoxia Assessment data
showing that watersheds containing less than 5%
cropped land contribute 235 970 metric tons N of
the 1 567 900 metric tons N delivered by the
MARB to the Gulf of Mexico. And these very
low cropped land density watersheds have relatively low inputs of non-agricultural anthropogenic N.
Average inputs of industrial plus municipal N are
39% that of the MARB average, and average input
of atmospheric N deposition is 68% that of the
MARB average (Goolsby et al., 1999, Tables 2.3,
4.3, 5.2, 5.7). These data indicate that a large
quantity of mineralized soil N emanates from soils
of the two thirds of the MARB which the Hypoxia
Assessment’s N budget assumed to have zero net
soil N mineralization.
Third, ‘the mass of organic matter (in kgykm2)
in the upper 30 centimeters (cm) of soil was
calculated as the product of the soil bulk density,
percent organic matter, and volume. The soil N
content was estimated as 3% of the organic matter
(Stevenson, 1994). The soil organic N was estimated to mineralize at a wnetx rate of 2% per year
in cultivated soils.... Total potentially wnetx mineralizable N estimates were computed for STATSGO
wState Soil Geographic Data Basex map units and
then generalized to counties using area-weighted
averages’ (Goolsby et al., 1999, p. 48). Stevenson
(1994, p. 7) reported the N content of SOM to be
approximately 5.8%, not 3% as cited by Goolsby
et al. (1999, pp. 48, 64). Troeh and Thompson
(1993, p. 98), whom Goolsby et al. (1999) cite
extensively, report that SOM is, on average, 5%
N. On the basis of this range in reported SOM N
values, the net mineralized soil N yry1 values for
the upper 30 cm of MARB cropped land range
from 9.5 to 13.1 million metric tons N yry1. The
given estimate of 6.8 million metric tons N yry1
E.C. Krug, D. Winstanley / The Science of the Total Environment 293 (2002) 1–29
(Goolsby et al., 1999, Table 5.2) is low by 2.7 to
6.3 million metric tons N yry1. This underestimate
is also significant, being larger than the reported
1.6 million metric tons N yry1 load from the
MARB to the Gulf of Mexico (Goolsby et al.,
1999, Table 5.7).
Fourth, the Hypoxia Assessment excluded net
soil-N mineralization below 0.3 m (1 foot) soil
depth. However, the soils of the MARB’s bread
basket are appreciably deeper than 0.3 m. These
cropped MARB soils have well above average
amounts of SOM and N to depths of 1 m, and
below (Marbut, 1929; Schreiner and Brown, 1938;
Jenny, 1941; Stevenson, 1959a; Geis et al., 1970;
Soil Survey Staff, 1975; Odell et al., 1984; Stevenson and Cole, 1999, pp. 46–56, 139–145). For
example, Stevenson (1982b, Table 3) reported
SOM data for 15 and 100 cm for prairie soils of
the North Central States — the mainstay soils of
MARB cropped land — and Marbut (1929) reported values for these soils to at least 150 cm depth.
Assuming that the soil N values of these soils are
the same in the top 30-cm of soil as they are in
the top 15-cm, then they contain 56% of their
SOM and N between 30 and 100 cm, and 65%
between 30 and 150 cm.
Prairie soils are deep and rich in SOM in part
because the prairie was dominated by prolific,
deep-rooted plant species. Depending upon soil
conditions and species, the dominant species of
tallgrass prairies sent roots down 1.5 to 4 m
(Weaver, 1926, pp. 95–97) and the average mass
of prairie roots was an order of magnitude greater
than the average root mass of annual crops (Reeder
et al., 2001, p. 142). Nevertheless, the major crops
of the MARB — corn, soybeans, wheat — have
extensive root systems which extend ;1 m, or
deeper, down into the soil (Weaver, 1926; Fehrenbacher and Snider, 1954; Fehrenbacher et al., 1969;
Mayaki et al., 1976; Emery, 1980; Goss, 1991;
Majchrzak et al., 2001). A typical cornfield has
182 000 km of roots per hectare — enough to
circle the world more than 4 times (Barber, 1992,
p. 24). In prairie soils corn roots regularly penetrate to 2 m (Kramer and Boyer, 1995, p. 126). In
Illinois, corn roots typically extend down 6 foot
(1.8 m) (Boone et al., 1978, p. 32), wheat 4 foot
(1.2 m) (Majchrzak et al., 2001), and soybeans
9
3–4.5 foot (0.9–1.4 m) (Fehrenbacher et al.,
1969); soybeans in Illinois extend their roots 1.2–
1.6 m under drought stress (Emery, 1980). Therefore, since SOM and its associated N is
mineralizable throughout the soil profile to depths
of 1 m and greater (Stevenson, 1959a, 1982b, p.
115; Brady, 1974, pp. 154–163; Cassman and
Munns, 1980; Powers, 1980; Power et al., 1985;
Baethgen and Alley, 1986; Hadas et al., 1986a,b,
1989; Turco and Sadowsky, 1995; David et al.,
1996, 1997), there is extensive net soil-N mineralization below 30 cm soil depth and this mineralized N is available for both uptake by crops and
leaching loss to the hydrosphere (Mengel and
Scherer, 1981; Baethgen and Alley, 1986; Kuhlmann et al., 1989; Li et al., 1990; Soon, 1998;
Mitsch et al., 1999, pp. 4–6). Such net soil-N
mineralization is excluded from the Hypoxia
Assessment’s MARB N budget.
Yet another reason for concluding that the
Hypoxia Assessment grossly underestimated net
mineralization of soil N in the MARB is extensive
drainage of the soils. An important reason why the
crop root systems typically grow so deep and crop
productivity is so high in the MARB breadbasket
is because of the drainage of extensive areas of
N-rich wetland soil into croplands (Alexander,
1905; King, 1918; Van Vlack and Norton, 1944;
Hewes, 1951; Pavelis, 1987); e.g. ‘Without drainage, it is hard to imagine the US Midwest as it
has developed to be the most productive agricultural area anywhere in the world’ (Fausey, 1993,
p. 519). During the last 200 years more than 80%
of the exceptionally nutrient- and N-rich wetland
soils of key agricultural North Central States have
been drained (CENR, 2000, p. 14). Overall,
approximately one quarter of the cropped land in
the MARB is drained (Pavelis, 1987), with
approximately 60% of the cropped-land drainage
concentrated in just five states — Illinois, Indiana,
Ohio, Iowa and Minnesota (Pavelis, 1987). The
Hypoxia Assessment recognizes that watersheds
yielding higher concentrations of NO3 –N are characterized by the presence of wet, high SOM soils
that have been drained by tile drains and ditches
to 1–1.5 m depth for optimum crop production
(Goolsby et al., 1999; Mitsch et al., 1999). The
Hypoxia Assessment also recognizes that drainage,
10
E.C. Krug, D. Winstanley / The Science of the Total Environment 293 (2002) 1–29
in and of itself, results in the oxidation of the vast
amounts of previously waterlogged subsurface
SOM. Such drainage results in the mineralization
of the vast amounts of soil N and conversion of
much of this mineralized subsoil N to NO3 –N
(Mitsch et al., 1999, pp. 4–6). Nevertheless, the
MARB N budget developed by the Hypoxia
Assessment does not recognize mineralization of
soil N as occurring below 30 cm soil depth
(Goolsby et al., 1999).
Not recognizing net soil-N mineralization below
30 cm is an especially serious error for drained
soils because the quantity of mineralizable soil N
increases with increasing wetness (Schreiner and
Brown, 1938; Jenny, 1941; Joffe, 1949; Soil Survey Staff, 1975; Troeh and Thompson, 1993). The
amount of mineralizable soil N increases from
6500 kg N hay1 to 218 000 kg N hay1 in the top
1 m of Illinois soils with increasing wetness
(Hopkins, 1910, pp. 77–87). These same soils are
common throughout the North Central States
(Hopkins, 1910, pp. 77–87; Bear 1929, p. 55;
Schreiner and Brown, 1938, p. 370) — the states
of greatest concern regarding N release from
cropped lands to the waters of the MARB (Goolsby et al., 1999; CENR, 2000).
The above points illustrate that the goodness of
fit between inputs and outputs and the achievement
of mass balance for the MARB which were used
to validate the current model and its fundamental
premise of N saturation are fortuitous and
erroneous.
At the state level, David and Gentry (2000)
applied a method for Illinois similar to that
employed by Jordan and Weller (1996) and
Howarth et al. (1996) by comparing anthropogenic
N inputs with riverine exports of total N from all
sources, anthropogenic and natural. This approach
assumes no natural N inputs and no change in
Illinois’ soil N reservoir. Such an approach forces
the conclusion that anthropogenic inputs (the only
inputs by definition allowed to exist) are the cause
of the outputs from all sources and causes.
Using this method, David and Gentry (2000)
concluded that since 1945, N fertilizer has provided ;60% of the total annual average net N input
of 32 kg N hay1 yry1 for 1980 to 1997 for Illinois
of whose 146 000 km2 area, 93 000 km2 is row
crop agriculture, principally corn-soybean rotation.
17 kg N hay1 yry1 of the 32 kg N hay1 yry1 of
net anthropogenic input was reported as showing
up as river N output; the 15 N hay1 yry1 difference
was called ‘surplus’ N. David and Gentry cited
the Drinkwater et al. (1998) 15-year corn-soybean
rotation study to justify the assumption that Illinois’ soil N reservoir remained unchanged. Using
arbitrary values of terrestrial and aquatic denitrification (10% of fertilizer input and 35% of river N
load, respectively) a ‘good fit’ of inputs to outputs
was achieved, balancing the state of Illinois’ N
budget to within 2%. In spite of the validation this
mass balance of inputs and outputs achieved,
analysis of the approach does not support the
conclusion that ‘This approach can be used as awnx
heuristic tool in identifying gaps in our knowledge
concerning the fate of N...’ (David and Gentry,
2000, p. 495) because both the input-output
approach and its budget-balancing ‘validation’ are
logically flawed. The approach that landscapes are
imperfect filters for anthropogenic additions of N
cannot account for the natural accumulation of N
in soils and waters. It does not address the fact
that approximately 40% of the original N content
of Illinois soils has been lost (Hoeft and Peck,
2000, p. 94); nor does it address the continued
loss of soil N by erosion and oxidation (mineralization) of SOM in Illinois (Hoeft and Peck, 2000,
p. 94) and the MARB. Indeed, the Drinkwater et
al. (1998) corn-soybean rotation study asserted to
show ‘no significant changes in soil N pools of a
conventional corn and soybean rotation during a
15-year period’ (David and Gentry, 2000, p. 504)
actually reported a statistically significant soil N
loss of 500 kg N hay1 — a loss of 33.3 kg N
hay1 yry1 — under conventional corn-soybean
rotation (Drinkwater et al., 1998, pp. 263–264).
On a local scale, David et al. (1996, 1997)
conducted an N mass balance for a 48 000 ha
headwater of the ERW in central Illinois — a
predominantly row crop (corn and soybean), tiledrained agricultural watershed. Unlike the subsequent statewide approach (David and Gentry,
2000), David et al. (1996, 1997) recognized that
a large pool of soil organic N exists in the
watershed, e.g., 11 000 kg N hay1 in the top 1 m
of Drummer soil: the predominant soil of the ERW
E.C. Krug, D. Winstanley / The Science of the Total Environment 293 (2002) 1–29
(David et al., 1996, p. 227) and 400 000 metric
tons N in the top 1 m of ERW soil (David et al.,
1997, p. 1047). It was concluded that, even if N
fertilization were eliminated, the overall disturbance from agricultural production would still lead
to high N concentrations and export. However, the
authors did not quantify such export (David et al.,
1996, 1997).
Generally, the analysis and conclusions of the
ERW were similar to those of the Hypoxia Assessment for the MARB. As with the Hypoxia Assessment (Goolsby et al., 1999), ERW net soil-N
mineralization was treated as an external input to
the watershed — the estimated net soil-N mineralization of 4696 metric tons N yry1 being approximately equal to the input of 4421 metric tons N
yry1 of chemical-N fertilizer (David et al., 1997).
As with the Hypoxia Assessment (Goolsby et al.,
1999), and consistent with the current model
results on the global scale, it was concluded that
external inputs to the ERW exceed terrestrial
outputs by 3671 metric tons N yry1, or 76.2 kg N
hay1 yry1 (David et al., 1997). As with the
Hypoxia Assessment (Goolsby et al., 1999), a
portion of this putative ERW N budget excess —
1688 metric tons N yry1 or 35.0 kg N hay1 yry1
— was reported to have been exported from the
watershed by the Embarras River (David et al.,
1997, pp. 1045–1046). The remaining unexplained
excess — 1983 metric tons N yry1 (41.2 kg N
hay1 yry1) — was assumed to be due to various
factors, including net accumulation of N in soil,
loss of N to the atmosphere by denitrification,
andyor due to erroneously high estimation of net
soil-N mineralization in the ERW (David et al.,
1997, p. 1047). As with the Hypoxia Assessment
(Goolsby et al., 1999), if net N mineralization is
not treated as an external input then the ERW
terrestrial N budget runs a deficit — y1025 metric
tons N yry1 (y21.3 kg N hay1 yry1). Add in
mineralized N loss to river export, and the ERW
N budget is y2713 metric tons N yry1 (y56.3
kg N hay1 yry1) in deficit, exclusive of consideration of denitrification and other volatilization
losses to the atmosphere and loss of organic N in
erosion and in solution.
Soils data indicate that the soil N reservoir of
the ERW has decreased considerably during the
11
20th century. Although David et al. (1996, p. 227)
report that the predominant soil type of the ERW
has 11 000 kg N hay1 in the top 1 m, turn-of-thecentury data show that it had 17 000 kg N hay1
in the top 1 m (Hopkins, 1910, pp. 77–87;
Hopkins et al., 1918). Futhermore, David et al.
(1997, p. 1047) report that there is 400 000 metric
tons N in the top 1 m of ERW soil, turn-of-thecentury data show that there was between 780 000
and 970 000 metric tons N in the top 1 m of ERW
soil (Hopkins, 1910, pp.77–87; Hopkins et al.,
1918). And, as with the MARB, by correctly
treating net soil-N mineralization as mineralization
of N already within the system, the ERW N budget
indicates that the ERW soil-N reservoir is still
being depleted.
Indeed, all of the above data indicate that the
soils of the contiguous continental USA are still
losing appreciable N in addition to the 1.75 billion
metric tons estimated already to have been lost by
the middle of the 20th century (Viets and Hageman, 1971, p. 14). In addition the above data do
not include the estimated 6.4 million metric tons
N yry1 being lost by erosion from the soils of the
contiguous continental USA to surface waters
(Gianessi and Peskin, 1981, p. 804). This erosion
N loss estimate does not include erosion losses in
snowmelt, from federal lands, or from irrigated
lands (Gianessi et al., 1980, p. 25).
In conclusion, the research findings of the current model invariably conform to the N-saturation
hypothesis. The current model is characterized by
partial and confused analyses and does not support
the development of a consistent and comprehensive analytical framework on any geographic or
temporal scale. The current model ignores much
data and the data it does recognize do not support
the conclusions drawn from them. In general terms,
the current model does not reconcile itself with
the fields of soil conservation and sustainable
agriculture. Namely, the current model does not
reconcile itself to concurrent environmental concerns about terrestrial N depletion, and related
issues of climate change; e.g. the loss of soil C
from anthropogenically-induced loss of SOM
which has increased atmospheric CO2 content
(Stuiver, 1978; Houghton et al., 1983; White,
1990; Harrison et al., 1993) and the desirability of
12
E.C. Krug, D. Winstanley / The Science of the Total Environment 293 (2002) 1–29
restoring the lost soil organic matter to offset the
greenhouse effect, as is reflected in calls for
increasing soil C sequestration through conservation tillage and other management techniques (e.g.
Barnwell et al., 1992; Li et al., 1994; Lal et al.,
1998). The current model, by not scientifically
accounting for both the addition and subtraction
of N from the environment by anthropogenic
activities, does not allow itself to conclude from
the available data that the net effect of anthropogenic activities over the last 200 years has been to
decrease the amount of reactive N in the terrestrial
biosphere on global, regional, and local scales.
4.2. A previous conceptual model
In this section a previous model is often
described in reference to the current one. This is
because in order to successfully describe something unknown, it must be described in terms of
what is known. With this previous model, scientists
operated under a more comprehensive conceptual
framework of N cycling, which acknowledged
explicitly that most reactive inorganic N is stored
and cycled in the terrestrial N reservoir (Viets and
Hageman, 1971; Rosswall, 1976; Stevenson,
1986). It was recognized that human activities and
natural agencies both add N to and remove N from
the terrestrial N reservoir and that these additions
and removals influence cycling within and beyond
this reservoir and its various pools. This previous
conceptual framework has been largely lost. Much
research conducted under the previous conceptual
framework also has been largely ignored, but much
of this research is relevant to contemporary concerns. For example, it was estimated that under
this previous conceptual framework some 100 000
experiments were performed on soil NO3 –N leaching alone (Viets and Hageman, 1971, p. 19).
The previous model had a framework comprehensive and consistent enough to incorporate the
then-extant data and concepts. For example, the
previous model was able to accommodate an
estimate of the net biospheric N cycle: 10 000
million metric tons reactive inorganic N yry1 was
the global estimate of which 5800 million metric
tons of reactive inorganic N yry1 were estimated
to cycle in the top 1 m of soil alone (Viets and
Hageman, 1971; Rosswall, 1976).
Furthermore, the previous model provides a
framework for the incorporation of data and concepts developed under the tenure of the current
model, data and concepts the current model does
not incorporate. Using the terrestrial N cycle as an
example, since the demise of the previous model,
numerous studies have been conducted to improve
estimates of net N mineralization in soils. Review
of this recent research shows that the old estimates
of net soil N mineralization rates were low (Vogt
et al., 1991; Abril et al., 2001). Also, much
research has since been conducted on gross N
mineralization (e.g. Schimel, 1986; Smith et al.,
1994; Takahashi, 2001; Wang et al., 2001). This
research enables old estimates of net terrestrial N
mineralization to be adjusted upward and these
improved net estimates to be adjusted upward to
gross N mineralization estimates. Additionally, it
is now known that the terrestrial N cycle is
appreciable below 1 m (Fliermans and Balkwill,
1989; Korom, 1992; Madsen and Ghiorse, 1993;
Simpkins and Parkin, 1993; Parkin and Simpkins,
1995; Chu and Alvarez-Cohen, 1998; Turco and
Sadowsky, 1995; Petsch et al., 2001). For example,
Madsen and Ghiorse (1993, p. 171) report in their
review of groundwater microbiology that, ‘..assuming that the biomass per kilogram of subsurface
sediments is 1000-fold less on average than that
found in the top 1 m of agricultural soil, then the
total biomass in subsurface habitats may be 40
times greater than that in the top 1 m of soil.’
While not incorporated in the current model, under
the previous model such new knowledge can be
incorporated to adjust upward the estimates of the
global and terrestrial N cycles — estimates against
which inputs of anthropogenic N may be
compared.
The previous model had a more comprehensive
and inclusive view of natural terrestrial N fixation
than the current model. Whereas under the new
model estimates of anthropogenic N fixation invariably grow with new information, estimates of
natural terrestrial N fixation do not (Kinzig and
Socolow, 1994; Galloway et al., 1995; Jordan and
Weller, 1996; Galloway, 1998; National Research
Council, 2000). Stevenson, (1986, pp. 116–131)
E.C. Krug, D. Winstanley / The Science of the Total Environment 293 (2002) 1–29
observed that natural terrestrial N fixation in soils
was estimated to be several times larger previously
than it is now. Stevenson also noted that the
number of known species of N-fixing bacteria had
greatly increased in recent years and that this
should be significantly increasing our estimate of
terrestrial N fixation (Stevenson, 1986, p. 123).
The previous model recognized that N fixation
is carried on in a surprisingly wide range of
conditions by a surprisingly wide range of organisms in terrestrial environments. Research conducted under the reign of the previous model
recognized that N fixation is conducted by Nfixing heterotrophic bacteria living in the soil as
well as by heterotrophic bacteria living in the guts
of termites, bark beetles, and aphids as well as in
the guts of earthworms, insects, snails and other
soil animals which feed on decaying organic matter, plus in the guts of protozoans, in the intestines
and rumen of animals, in their feces after excretion
and in association with fungi as well as by bluegreen algae (Allison et al., 1937; Csaky and Toth,
1948; Bergersen and Hipsley, 1970; Granhall and
Ciszuk, 1971; Breznak et al., 1973; Hobson et al.,
1973; Cornaby and Waide, 1973; Sharp, 1975;
Baines and Millbank, 1978). For example, humans
living on a principally vegetable diet have been
shown to have intestinal N2-fixation approximately
equal to their dietary intake of N (Bergersen and
Hipsley, 1970). In unsettled areas with appreciable
termite activity, nitrification of N fixed in the guts
of termites raises the NO3 –N concentrations of
whole aquifers to concentrations appreciably over
the health standard of 10 mg NO3 –N ly1 presenting a health hazard to human settlement (Barnes
et al., 1992). Regarding soil micro fauna, ‘The
results reported here confirm previous observation...that the presence of N2-fixing micro-organisms in the gastro-enteric cavity is a common
feature of soil-inhabiting fauna which feed on
decaying organic matter....The weight of the soil
micro fauna can total up to 1 tonne hay1 or
more...and soil organic matter has been estimated
to pass through the fauna intestine several times
per year, on average...’ (Citernesi et al., 1977, pp.
71–72). Assuming that soil micro fauna process
only the 192.8 metric tons SOM hay1 in ERW
topsoil once per year, assuming that SOMs1.724
13
Soil Organic Carbon (SOC), and assuming that 1
atom of N is fixed per 10 000 atoms of SOC
processed, the N-fixing intestinal flora of soil
micro fauna fix 13.05 kg N hay1 yry1. For the
entire headwaters of the ERW, this comes to ;600
metric tons N yry1. This estimate of N fixed by
the intestinal flora of soil micro fauna in the
headwaters of the ERW can be compared to the
140 metric tons N yry1 estimate for atmospheric
N deposition to the watershed (David et al., 1997,
p. 1045).
Under the previous model, blue-green algae —
commonly thought of as autotrophic organisms
living mostly in water — were also recognized as
living in and fixing appreciable amounts of N in
moist, well-drained, and even excessively drained,
very dry soils (Fogg, 1956; Mayland and McIntosh, 1966; Stewart, 1969; Clark and Paul, 1970).
Under the previous model, free-living, blue-green
algae were recognized as an important N-fixing
component of natural and agricultural soils; e.g.
‘Practically all cultivated soils and grasslands have
a rather abundant flora of blue-green algae, various
species of Nostoc wN-fixing blue-green algaex
being especially common. In fact the majority of
the members of this genus are terrestrial.... Calculations of the actual mass of algal cells in soil
show that their weight is often almost as great as
that of the bacteria.’ (Allison et al., 1937, p. 457).
These N-fixing, blue-green algae can live in the
light or in the dark. These algae have long been
known to thrive on CO2, andyor carbohydrates
and humic acids (e.g. Allison et al., 1937; Fogg,
1949; Fay, 1965; Watanabe and Yamamoto, 1967;
Khoja and Whitton, 1971). These realities obliterate the nice neat lines often drawn between the
plant (autotrophic) and animal (heterotrophic)
kingdoms. For example, six years of study of the
common soil alga, Nostoc muscorum showed,
‘Nostoc muscorum will live and grow for months
in the dark and form normal appearing chlorophyll
if given a suitable energy source, such as glucose.
Growth and nitrogen fixation take place rather
slowly but the quantities fixed per unit energy are
usually large....An average fixation of 9.6 mg. N
per gram glucose consumed, with a maximum
value of 14.8 mg. N under ordinary atmospheric
conditions of growth, is higher than commonly
14
E.C. Krug, D. Winstanley / The Science of the Total Environment 293 (2002) 1–29
obtained with Azotobacter or other nitrogen-fixing
bacteria’ (Allison et al., 1937, pp. 455–456). Bluegreen algae have been shown to be a potentially
significant source of N to temperate agricultural
soils. Experiments with raw soil samples (Granhall, 1975) and soils of wheat fields (Witty, 1979;
Witty et al., 1979) show a complex relationship
between N fixation by blue-green algae, N fertilization, and environmental conditions. Increasing
moisture, temperature, and pH generally enhance
N fixation (Granhall, 1975; Witty, 1979; Witty et
al., 1979). N fertilizers have a mixed effect on N
fixation by blue-green algae in soils. Initially N
fertilization depresses N fixation. However, later
in the season N fertilizer addition can enhance N
fixation by increasing canopy cover (reducing
dessication of surface soil) and increasing organic
matter production. On wheat fields receiving 48
kg N hay1, soil blue-green algae were estimated
to fix 25–28 kg N hay1 prior to harvest. Unfertilized control wheat fields fixed 13–19 kg N hay1
prior to harvest (Witty et al., 1979). In wheat field
inoculation experiments using the blue-green
algae, Nostoc ellipsosporum, algal N fixation rates
were elevated by spring application of 80 kg N
fertilizer hay1 (Witty, 1979). Spring application
of 196 kg N hay1 suppressed algal N fixation until
so late in the 4 month growing season that 1.4 kg
N hay1 was fixed prior to harvest (Witty et al.,
1979). In light of the conventional consensus belief
that no more than 6 kg N hay1 yry1 is added to
soils in the USA by the combined activities of
asymbiotic N-fixing microorganisms (Stevenson
and Cole, 1999, p. 154), blue-green algae living
in and on soils can fix impressive amounts of N,
even when such N fixation is hindered by high
rates of chemical-N fertilization.
As already mentioned, the previous model recognized that N fixation occurs under a wide range
of conditions by a wide range of organisms in
terrestrial environments. N-fixing heterotrophic
bacteria are common among those decomposing
organic matter in soils. And research in recent
decades using long-established methodologies has
expanded the range of N fixation by showing that
asymbiotic heterotrophic N-fixing bacteria and
algae are even more highly concentrated on and
in the organic-rich root, stem, and leaf surfaces of
plants, including corn, wheat, and even soybeans,
than they are in bulk soil (Stewart, 1969; Ruinen,
1970, 1974; Seidler et al., 1972; Bessems, 1973;
Aho et al., 1974; Dobereiner, 1977; Parker, 1977;
Purchase, 1977; Finke and Seeley, 1978; Kana and
Tjepkema, 1978; Moeller and Roskoski, 1978;
Ogan, 1979; Watanabe et al., 1979; Schink et al.,
1981; Lending, 1984; Ueda et al., 1995; Kapulnik,
1996).
New methodologies have further expanded the
range and quantity of natural N fixation. Many, if
not most, N-fixing bacteria are not identifiable by
traditional means (Paerl, 1985; Ueda et al., 1995;
Zehr et al., 1995) — the means of identification
available during the reign of the previous model.
However, since then, DNA isolation, identification,
and amplification techniques have been developed.
These techniques have opened up a new world of
N-fixing bacteria (e.g. Ueda et al., 1995; Zehr et
al., 1995, 1998; Carpenter et al., 1999). With this
new technology and an explosion of new research,
N fixation has been found to be going on under
previously unknown conditions and at previously
unknown rates. The new findings show that natural
terrestrial and global N fixation and cycling are
appreciably underestimated (Bridges, 1981; Schink
et al., 1981; Lending, 1984; Zehr et al., 1995,
1998, 2001; Ueda et al., 1995; Bushaw et al.,
1996; Kapulnik, 1996; Eviner and Chapin, 1997;
Karl et al., 1997; Gruber and Sarmiento, 1997;
Stark and Hart, 1997; Bagwell et al., 1998; Chu
and Alvarez-Cohen, 1998; Currin and Paerl, 1998;
Curtis and Waller, 1998; Wei and Kimmins, 1998;
Braun et al., 1999; Carpenter et al., 1999; Clawson
and Benson, 1999; Hendricks and Boring, 1999;
Sachs and Repeta, 1999; Stepanauskas et al., 1999;
Guildford and Hecky, 2000; Hood et al., 2000;
Boudreau et al., 2001). The previous model can
readily incorporate these new findings and encourage additional research on natural N fixation.
5. Comparison of the current and previous
conceptual models
The dominant conceptual model has changed in
recent decades. Fewer concepts and data are
included and those remaining are not treated consistently and comprehensively — especially in
E.C. Krug, D. Winstanley / The Science of the Total Environment 293 (2002) 1–29
regard to the supply and source of reactive N to
the biosphere, the treatment of N cycling within
the terrestrial N reservoir and the quantification of
N fluxes to and from this reservoir.
The recent regional (Goolsby et al., 1999;
CENR, 2000), state (David and Gentry, 2000),
and local (David et al., 1996, 1997) studies of N
cycling and N mass balances have some significant
differences and inconsistencies, but they share a
broadly similar conceptual framework consistent
with the current global model and its N-saturation
hypothesis (Galloway et al., 1995). They all focus
on increasing N input since ;1950, especially
increasing use of chemical-N fertilizer. On all
spatial and temporal scales, the reported results are
consistent with the idea that landscapes have
achieved N saturation and are leaking excess N
input from the N-saturated soil reservoir to the
hydrosphere. A key mechanism by which the
current model grows soil N reservoirs to the point
of N saturation is by treating mineralization of N
already in the soil reservoir as an external N input
to the soil N reservoir while at the same time
acknowledging mineralization of the soil N reservoir as the mineralization of N already in the
system. Not only is the current model internally
inconsistent in this respect, it also is inconsistent
with the previous model and with current environmental concerns of soil conservation and global
warming (Stuiver, 1978; Houghton et al., 1983;
White, 1990; Barnwell et al., 1992; Harrison et
al., 1993; Li et al., 1994; Lal et al., 1998).
The current model, focused on the concept of
recent anthropogenic N saturation of the terrestrial
N cycle, remains silent on many observed phenomena and measured data, which were addressed by
the previous model.
First, the current model is thus inconsistent with
the previous model which recognized that erosion
of soil organic N was the most important source
of N delivery to surface waters from agricultural
lands in the USA, even though improved conservation has greatly decreased erosion losses of N
to the hydrosphere in recent decades (e.g. Stevenson, 1986, p. 148). Indeed, the Hypoxia Assessment is internally inconsistent as it states that soil
erosion is still the most important source of N loss
from cropped land to waters in the MARB (Doer-
15
ing et al., 1999, pp. 27–30, Table 4.1–1) but does
not explicitly include loss of soil N to erosion in
its N cycle or N mass budget (Goolsby et al.,
1999; CENR, 2000).
Second, unlike the previous model, the current
model does not reconcile itself with the fact that
the load of suspended sediment carried by the
MARB to the northern Gulf of Mexico has
decreased by an estimated 70% from background
levels — 50% since the 1950s (Keown et al.,
1986; Kesel, 1988). Assuming that the composition of MARB suspended sediment (Malcolm and
Durham, 1976, p. F12) has remained constant over
time, such a decrease in suspended sediment equals
a decrease of 1.9 million metric tons N yry1 from
the MARB to the northern Gulf of Mexico (Keown
et al., 1986; Kesel, 1988). This decrease of 1.9
million metric tons N yry1 can be compared to
the recent total loading rate of 1.6 million metric
tons N yry1 (Goolsby et al., 1999, Table 5.7),
which is widely asserted to represent a severalfold increase of N loading from the MARB to the
northern Gulf of Mexico over background levels.
Third, unlike the previous model, the current
model cannot adequately explain the cases of large
concentrations of N in ground- and surface waters
prior to the 1950s, the high rates of loss of NO3 –
N in waters routinely reported to be issuing from
unfertilized soils prior to the 1950s, reported cases
of high natural NO3 –N, or more recent studies
showing high rates of loss of NO3 –N from unfertilized soils (e.g. Jolley and Pierre, 1977; Robbins
and Carter, 1980; Owens, 1990; Johnston et al.,
1994; Addiscott, 1996; Randall et al., 1997).
In Illinois, the average concentration of total N
in the lower Illinois River from 1894 to 1899 was
3.7 mg N ly1 (Palmer, 1903; Krug and Winstanley,
2000). In its major tributaries — the Kankakee,
Des Plaines, Fox, Vermilion, Mackinaw and Spoon
Rivers — the average concentration of total N in
1921–1922 peaked in December and was 6.7 mg
N ly1 (Hoskins et al., 1927; Krug and Winstanley,
2000). Undoubtedly, human and animal wastes
augmented the natural N.
During the 1940s in Illinois, nearly 6000 well
samples from private water supplies of all types
from all sections of the state showed that 20% had
)10 mg NO3 –N ly1. More than 40% of samples
16
E.C. Krug, D. Winstanley / The Science of the Total Environment 293 (2002) 1–29
in central Illinois counties had )10 mg NOy
3 –N
ly1, and more than 20% had )100 mg NO3 –N
ly1. These water supplies were reported to show
no correlation with animal andyor human wastes
(Weart, 1948). In 1970, Illinois State Water Survey
scientists acknowledged the earlier high NO3 –N
groundwater values and added, ‘the records show
this happened long before commercial, N fertilizer
usage became significant’ (Harmeson and Larson,
1970, p. 33).
In Iowa, a measured decrease of NO3 –N concentrations has been reported in groundwaters
under cropped land in the face of increasing
fertilizer usage (e.g. Eidem et al., 1999). These
surface water and groundwater findings on N
chemistry are consistent with soil-N research conducted in Illinois prior to the use of chemical-N
fertilizers, ‘Large amounts of nitrogen are annually
removed in crops and carried away in drainage
water.... drainage water may carry away 100
pounds or more w112 kg NO3 –N hay1yry1 or
morex’ (Linsley, 1931, p. 4). Prior to the use of
chemical-N fertilizer, large amounts of NO3 –N
were lost from unfertilized soil: on a yearly- and
even on a rain-event basis. For example, Whiting
and Schoonover, (1920, pp. 43–44) report that
with 3 in (7.6 cm) of rain occurring between
March 10 and March 22, 1917, the plow layer
(top 6.67 in, 17 cm) of soil of unfertilized corn
fields of the University’s Agricultural Experiment
Station lost 24.3 kg NO3 –N hay1. Assuming no
water losses to evapotranspiration and no N losses
to denitrification (counteracting effects), 7.6 cm
of precipitation produced 7.6 cm of runoff with an
average of 31.9 mg NO3 –N ly1. For soils fertilized
(receiving manure from stables), the plow layer
lost 33.9 kg NO3-N hay1; runoff is calculated as
having had 44.4 mg NO3 –N ly1. On average,
54.6% of the cornfields’ plow layer’s NO3 –N was
lost in 12 days. Whiting and Richmond (1921)
note that decomposition of clover (a legume)
produces large amounts of NO3 –N in a short time,
21 to 28 days, ‘From this year’s w1919x results on
the brown silt loam it is evident that the nitrate
production on this type of soil is greatly enhanced
by the plowing under of green sweet clover.
Another important fact is that approximately one
ton (water-free basis) of spring growth of sweet-
clover tops, together with the roots and fall residues, furnished as much nitrate as 19.8 tons of
average farm manure’ (Whiting and Richmond,
1921, p. 261). Whiting and Richmond (1921)
noted that this NO3 –N was extremely susceptible
to leaching. In 1935, the University of Illinois
started long-term studies of nutrient loss from
typical agricultural soils to surface water and
groundwater using triplicate, undisturbed, 1-m
deep, uncropped, unfertilized, soil erosion-type
lysimeters from 8 different types of Illinois prairie
soils (Stauffer, 1942). Over a 10-year period, the
average loss of soil NO3 –N to surface water and
groundwater from these 8 soils averaged 80.4 kg
N hay1yry1 (Stauffer and Rust, 1954).
Similar to Illinois, in the Canadian Prairie Provinces )10 mg NO3 –N ly1 are found in up to
20% of wells tested, and NO3 –N concentration
frequencies today are reported to be generally no
higher than the frequencies measured back to the
1940s (Harker et al., 1997, p. 44).
The results of a survey of well water in agricultural areas of central Canada (Ontario) also suggest that agricultural activity over the past 50
years, including the widespread and intensive use
of chemical-N fertilizers, has not significantly
changed the amount of NO3 –N added to groundwater (Fairchild et al., 2000, p. 65).
Also similar to this old Illinois groundwater data
are the 1930s and 1940s US Geological Survey
and Texas Board of Water Engineers groundwater
data for Texas. To quote the abstract of their
findings, ‘Ground water in many parts of Texas
contains nitrate in excess of 20 ppm (parts per
million) as nitrate. Approximately 3000 of the
20 000 nitrate determinations made of water from
wells in Texas showed more than 20 ppm of
nitrate. The public water supplies of 27 Texas
towns and cities contained more than 50 ppm of
nitrate.... Most of the high nitrate in ground water
is found in wells less than 200 foot deep and
mainly in water from late Tertiary and Quaternary
formations; however, high nitrate was found in
water from all kinds of rocks of all ages. The
presence of high nitrate in ground water appears
to be unrelated to rainfall, geography or cultivation’ (George and Hastings, 1951, p. 450). More
recently, the average concentration of NO3 –N in
E.C. Krug, D. Winstanley / The Science of the Total Environment 293 (2002) 1–29
233 wells sampled in Runnels County, Texas in
1969 was reported to be 64.8 mg N l-1 (Jones,
1973). Most water-well contamination was reported to be due to naturally occurring NO3 –N. The
few very high NO3 –N contamination problems
()225 mg NO3 –N ly1) were reported to be
principally due to pollution from barnyards. Only
27 of the analyses had less than the maximum
contaminant limit for NO3 –N (10 mg NO3 –N
ly1).
Thousands of wells sampled in Montana and
North Dakota gave similar results (Bauder et al.,
1993) to the data from Illinois, Texas and Canada.
Groundwater underlying virgin prairie in Colorado,
Nebraska and Alberta had concentrations of NO3 –
N exceeding the health limit of 10 mg NO3 –N
ly1 (Viets and Hageman, 1971, Table 11; Boyce
et al., 1976; Hendry et al., 1984; Krug and Winstanley, 2000, p. 26). Sediments and sedimentary
rocks can be natural sources of appreciable NO3 –
N (e.g. Stevenson, 1959b; Graf, 1960, p. 37; Chalk
and Keeney, 1971; Power et al., 1974). In the
essential absence of N fertilizer application, waters
seeping out of the Colorado Shale and Fort Union
geologic formations — which underlie much of
the Northern Great Plains — average 360 mg
NO3 –N ly1 and 135 mg NO3 –N ly1, respectively
(Brown et al., 1982). Water draining Upper Cretaceous to Tertiary sedimentary rocks in the agricultural San Joaquin Valley of California ‘contain
up to 2000 mgyliter indigenous nitrate N in the
soil solution...the geochemical effects of natural
N-containing systems often are overlooked when
explanations are sought for excessively high concentrations of nitrate in soils’ (Strathouse et al.,
1980, p. 54). Limestone formations underlie much
of southern Wisconsin and northern and central
Illinois. Analyses of limestone formations in this
region found ‘that many limestones are a potential
source of nitrate to percolate waters and that
geological contributions should be considered
when evaluating sources of nitrate to ground
waters’ (Chalk and Keeney, 1971, p. 42).
Furthermore, George and Hastings (1951) predicted the apparent future increase of NO3-N since
measured in groundwater, ‘Nitrate is probably
more prevalent in water from shallow wells
throughout the US than is generally known’ (p.
17
450) as analysts had shied away from analyzing
NO3 –N as well as shallow wells, instead preferring
to sample and analyze deeper, more productive
wells. However, as George and Hastings (1951)
noted, the increasing interest in sampling shallow
wells, and increasing interest in measuring NO3 –
N could, in and of itself, give the appearance of
increasing NO3 –N ‘contamination’ of groundwater. The subjectivity of observation problem is well
known to the scientific community. Using geology
as an example: ‘The value of the recent volcanological record is obvious to geologists.... The
limitations of the volcanological record are not as
obvious, however, and require continued emphasis
to caution anyone brash enough to mistake the
record for the reality’ (Simkin et al., 1981, p. 22).
Some of the illustrations Simkin et al. (1981) used
to demonstrate subjectivity were the record of
volcanic activity markedly jumping up after the
invention of the printing press, markedly jumping
down during World War I and World War II and
rebounding after the wars. Unfortunately, subjectivity of groundwater N measurement appears to
be a problem that has yet to be adequately
resolved.
Unlike the previous model, the current model
does not adequately explain many trends seen in
the N chemistry of surface waters. Examples can
be given from different parts of the world to
demonstrate that changes in N-fertilizer use are
not always associated with changes in N concentrations in rivers. In Norway, it was reported that
fertilizer use had decreased by 70% or more since
1989y1990, but by 1997 no significant changes
had been registered in the major rivers (Vagstad
et al., 1997, p. 272). In Iowa, Keeney and DeLuca,
(1993) demonstrated that there has been essentially
no increase in NO3 –N concentration in the Des
Moines River since 1945, even though N-fertilizers
were not widely used in Iowa until the 1950s after
which the N fertilization rate increased dramatically; e.g. 137 kg N hay1 yry1 was being applied
to corn by 1990. Concentrations of NO3 –N in
streams draining the Bluegrass region of Kentucky
have remained essentially unchanged (average
;5.5 mg NO3 –N ly1) between 1921 and 1990
(McHargue and Peter, 1921; Thomas and Crutchfield, 1974; Thomas et al., 1992) in spite of the
18
E.C. Krug, D. Winstanley / The Science of the Total Environment 293 (2002) 1–29
essential non-use of N fertilizer prior to the 1940s
(Thomas and Crutchfield, 1974, p. 49) with use
of N fertilizer increasing to 167 000 metric tons
yry1 by 1989 (Thomas et al., 1992, p. 147). In
the Ohio River, concentrations of NO3 –N have
changed very little since the 1950s, even though
N inputs to the watershed are large and N-fertilizer
use has increased (Goolsby et al., 1999) and the
cropped soils in the Ohio River Basin are largely
tile drained (Brezonik et al., p. 31). In Illinois, the
increase in NO3 –N concentration in the Lower
Illinois River from ;1950 to ;1970 was asserted
to be causally connected to increased N-fertilizer
use (Commoner, 1970), but NO3 –N concentration
decreased after 1970 in the face of continued
increase in N-fertilizer use (Krug and Winstanley,
2000, Fig. 27).
Whereas such phenomena are not adequately
explained by the current model, the previous model
did; e.g. ‘Many fertilized fields are contributing
less nitrate nitrogen to our streams and undergroundwater supplies than they did many years
ago. This is because the total nitrate nitrogen
currently available from the soil organic matter
and fertilizer is less than the soil organic matter
alone furnished at the beginning of cultivation’
(Stewart, 1970, p. 58). And, ‘By 1969, the annual
application of almost 7 million tons of nitrogen
was still insufficient to replace the yearly drop in
the soils capacity to supply this vital element in
plant growth’ (Library of Congress Environmental
Policy Division, 1973, p. 1297). These data are
consistent with the finding of the National
Research Council (1972) that the amount of N in
cropped land was declining.
Furthermore, the previous model has long recognized that natural rates of terrestrial N fixation
are appreciably greater than the current model
recognizes (e.g. Stevenson, 1986, pp. 116–131;
Stevenson and Cole, 1999, p. 154). For example,
looking at the early accounts of the MARB, the
early high NO3 –N concentrations in Kentucky
streams and rivers were stated to be due to naturally high rates of N fixation and nitrification in
soils developed from phosphorus-bearing limestone (McHargue and Peter, 1921). In traveling
through the Ohio Valley in the 1790s, Imlay, 1916
(1793, p. 11) noted that nitrate deposits were so
common in the predominantly limestone caves and
recesses of the region protected from flushing by
flowing water, that many of the settlers were using
it to make their own gunpowder. Late 20th century
studies of streams and rivers in Kentucky (Thomas
and Crutchfield, 1974; Thomas et al., 1992) concluded that concentrations of NO3 –N were correlated to geology (being high in phosphorus-bearing
limestone areas) and concentrations of NO3 –N did
not increase with increasing N-fertilizer use.
However, the current model does not include
much relevant science, data and observations, such
as the early observations of large amounts of
natural NO3 –N, high rates of natural N fixation,
the body of evolving scientific research which
shows that natural N fixation is greater and more
pervasive than even that realized under the previous model and the fact that soils are still losing
NO3 –N even when not fertilized. An important
reason for the inability of the current model to
readily explain many environmental trends, or the
lack of trends, is its inadequate treatment of terrestrial N.
The inadequate accounting of soil and geologic
N, natural N fixation, and net N mineralization in
the current model also results in overestimation of
the role of N fertilizer in riverine N fluxes. In
comparison with the above high fluxes of N from
soils on which no fertilizer was applied, average
total N riverine flux from the agricultural state of
Illinois in 1980–1997 was reported to be 17 kg N
hay1 yry1 (David and Gentry, 2000, p. 502). The
average N flux from the ERW, a tile-drained,
heavily fertilized area with primary cropping of
corn and soybean (David et al., 1997), was reported to be 24 kg N hay1 yry1 (David and Gentry,
2000, p. 502). N fertilizer is typically applied at a
rate of 130–190 kg hay1 yry1 on corn in central
Illinois. Depletion of the soil organic N reservoir
and the resulting transfer of large quantities of
soluble and particulate N to the hydrosphere were
either not explicitly recognized, or not quantified.
Recognizing that there are large quantities of
natural N in the soil, and that crops take up a
considerable quantity of N from this soil reservoir,
as well as from N fertilizer, the question arises as
to how much of the 17 kg N hay1 yry1 riverine
flux from Illinois and the 24 kg N hay1 yry1 flux
E.C. Krug, D. Winstanley / The Science of the Total Environment 293 (2002) 1–29
from the ERW is due to the output of ‘old’ soil
organic N and the mineralization of ‘old’ soil
organic N? Simply assuming it is zero (David and
Gentry, 2000) is not scientifically valid.
The more complete and realistic treatment of
soil organic N, as was characteristic in the previous
model, leads to further questioning of the validity
of the regional, state, and local N budgets cited
above. Had an annual net mineralization rate of
soil organic N of 1–3% in the top 1 m of the soils
of the headwaters of the ERW been used, this
would have yielded soil mineralization of 109–
327 kg N ha yry1,, compared with the range of
45–140 kg N hay1 used (David et al., 1996, p.
228; David et al., 1997, p. 1041). Had an annual
net mineralization rate of soil organic N of 2% in
the top 2 m in the ERW been used, this would
have yielded net soil mineralization much larger
than the reported value of 4696 metric tons N
yry1 (David et al., 1997, p. 1045). Had net soilN mineralization been included in the N budget of
Illinois (David and Gentry, 2000), and had a net
annual mineralization rate of soil organic N of 2%
in the top 2 m of the 93 000 km2 of row crops in
Illinois been used, assuming an average statewide
average soil N content of 7000 kg N hay1, the
resulting net generation of inorganic soil N would
not have been zero. It would have been approximately the same as all the anthropogenic N inputs
identified, ;1 300 000 metric tons N yry1 (David
and Gentry, 2000, p. 497). Consideration of total
(gross) N mineralization would have approximately doubled the ;1 300 000 metric tons soil N
yry1 estimate. Recognizing that there is a soil N
cycle in soils that are not cropped, and that there
is natural N fixation, would further increase the
estimated size of the soil N cycle to more than
two-fold greater than estimated anthropogenic N
inputs to Illinois. Regarding loss of soil N to
erosion, while David and Gentry (2000) assume
no erosion loss of N, the Hypoxia Assessment
(Doering et al., 1999, Table 4.1–1) reported that
soil erosion for conventional corn-soybean rotation
in the Corn Belt results in a loss of soil N of 44
kg N hay1 yry1, and under no-till corn-soybean
rotation the loss of soil N is 35 kg N hay1 yry1.
Applying these values to Illinois as a whole, and
assuming land not in row crops has no erosion,
19
soil N erosion loss comes to 22 to 28 kg N hay1
yry1 for Illinois, as compared to the reported loss
of 17 kg N hay1 yry1 loss in surface water (David
and Gentry, 2000, p. 502). Thus, not counting for
net soil N mineralization and soil N erosion losses
are serious omissions from the N budget developed
for Illinois (David and Gentry, 2000).
For the MARB, had a more valid estimate of
the N content of SOM in the cropped soils been
used, the estimated net N mineralization in the
MARB would have increased from 6.8 million
metric tons N yry1 (Goolsby et al., 1999, Table
5.2) to 9.5–13.1 million metric tons N yry1. Had
a net annual mineralization rate of soil organic N
of 2% in only the top meter in the MARB been
used, instead of the top 30 cm, this would have
approximately doubled the estimated net mineralization to ;19 to ;26 million metric tons N
yry1. Consideration of soil organic cycling in the
top 2 m would have further increased the estimate
of net soil mineralization — as would have consideration of net mineralization in all 300 million
ha of MARB rather than just the ;100 million ha
of cropped soil. Consideration of total (gross) N
mineralization would have approximately quadrupled the estimate of soil-N mineralization in
MARB cropped lands alone. Consideration of
natural N fixation would have raised it further.
Balancing the inputs and outputs of the MARB
was achieved by incorrectly estimating the amount
of soil-N mineralization, by treating the estimated
annual rate of net soil-N mineralization as an
external input to the MARB, by ignoring natural
N fixation, and, again, by not explicitly quantifying
the loss of soil N by erosion.
Overall, the current model does not provide a
consistent and comprehensive framework for evaluating the effects of human activities on the
terrestrial N cycle. The model does not treat
explicitly and fully the terrestrial cycling of ‘old’
N, and does not comprehensively treat the input
of naturally-fixed N. The current global model
recognizes explicitly only approximately 5%, or
less, of the anthropogenic plus net natural N cycled
annually on the continents and, thereby, overestimates by up to ;50-fold the effect of anthropogenic N additions in intensifying the global
terrestrial N cycle. Furthermore, the current model
20
E.C. Krug, D. Winstanley / The Science of the Total Environment 293 (2002) 1–29
forces the conclusion of N saturation by not
scientifically accounting for both the addition and
subtraction of N from the environment by anthropogenic activities whose net effect has been to
decrease the magnitude of the global terrestrial N
cycle.
The more comprehensive treatment of N cycling
provided by the previous model changes the magnitude of global N cycling on the continents and
the direction of agriculture’s calculated impact on
global, regional, and local terrestrial N cycling
over time. The previous model also is better able
than the current model to explain trends in N
concentrations and fluxes in many ground- and
surface waters, as will be further expounded upon
in a subsequent publication on the aquatic N cycle.
6. Conclusions and recommendations
Uncertainties in N cycling and N mass-balance
studies are large and many questions arise. For
example, the current global model reports that N
did not accumulate prior to the 20th century. If
this is true, then it must be asked how )300 000
kg N kmy2 accumulated in agricultural soils in
North America? And if natural ecosystems were
N limited and hydrologic losses of reactive N were
minimized, how could up to 30% of the N formed
by terrestrial fixation be lost to rivers and coastal
oceans? Why does the Mississippi-Atchafalaya
River today export only ;8.0% of total identified
N inputs to the MARB at a time when terrestrial
ecosystems are alleged to be more N saturated and
more leaky than in the pre-Industrial era?
N-cycling and N mass-balances studies on all
geographical and time scales are characterized by
high variability and uncertainty. From this, studies
such as the Hypoxia Assessment (e.g. Goolsby et
al., 1999, p. 67) have called for more research to
better understand the linkages between agricultural
and other human activities on the flux of N to the
hydrosphere. While this sounds hopefully progressive, it is not enough. Since the current model
refuses to falsify itself, using it is not the best way
to progress. Progress will be better served by the
implementation of a conceptual framework that
enables scientists to objectively process and assess
existent data and concepts, and allows us to design
research to generate additional data and concepts
that will broaden and deepen our understanding of
the N cycle. With this, some basic agronomic
information can be applied to improve N-cycling
and mass-balance studies. For example:
1. In the MARB and other parts of the world, the
amount of soil N has been seriously depleted
since the time of cultivation of virgin soil and
data show soil N is still undergoing depletion.
2. Crops such as corn, soybeans, and wheat extend
roots down 1–2 m into the soil and extract
water and nutrients from this depth and below.
3. In Illinois and other Midwest states approximately 20–30% of the crop area is drained by
tile systems at a depth of 1–1.5 m. It has been
long known that the potential sources of the N
content of the leachate include all the N above
the tiles and N drawn up from below the tiles
by lateral drainage (e.g. Pickels, 1941, p. 232;
Kirkham, 1958; Freeze, 1972; Abdul and Gillham, 1984; Waddington et al., 1993; Cirmo and
McDonnell, 1997).
4. There is a need to consider whole plant N
requirements in N mass-balance studies, not just
the N standing above-ground at harvest. For
example, corn-soybean crop rotation studies typically do not account for the N needed to grow
the leaves and roots of soybeans and the roots
of corn (e.g. Jordan and Weller, 1996; David et
al., 1997; Galloway, 1998; Goolsby et al., 1999;
Smil, 1999; CENR, 2000; David and Gentry,
2000; National Research Council, 2000). If soil
N in the top 2 m is not included in N massbalance studies, then it is impossible to conduct
a reasonable N mass balance; and it will be
impossible to find enough N to develop wholeplantysoil microbial N budgets (e.g. Rosswall,
1976) capable of producing the corn and soybean yields that are harvested.
5. It has been long known that legumes symbiotically fix atmospheric N and can, therefore, build
up SOM and N when legumes are not harvested.
When legumes are harvested, however, more
soil-derived N is removed in harvest than is
returned to the soil as symbiotically-derived N
(Russell, 1961, pp. 346–347). While it is popular to believe that growing soybeans in corn-
E.C. Krug, D. Winstanley / The Science of the Total Environment 293 (2002) 1–29
soybean rotation is a net soil N builder, for
soybeans in the Corn Belt, approximately onethird to two-thirds of the harvested N is soilderived N. Soybeans in corn-soybean rotations
are net soil N depleters (Heichel and Barnes,
1984; Hoeft and Peck, 2000, p. 97). In Illinois,
where approximately one-third of the N in
harvested soybean is soil derived, the net depletion of soil N by soybeans is calculated to
average 63 kg N hay1 yry1 based on the aboveground plant N budget (Hoeft and Peck, 2000,
p. 97).
Regarding more generalized analysis, we recommend that N studies adopt a comprehensive
and consistent treatment of the reservoirs and
cycling of terrestrial N. Such an approach should
include quantification at global, regional, and local
scales of (a) depletion of soil N from the time of
cultivation of virgin soils, (b) transfer of soluble
and particulate organic and inorganic N from the
geosphere to the hydrosphere, (c) mineralization
of soil organic N down to levels below the root
zone, (d) nitrification of fixed ammonia in the
geosphere, (e) natural N fixation, (f) quantification
of the role of precipitation in mineralizing soil
organic N and transferring N from the geosphere
to the hydrosphere, separate from external N inputs
and (g) quantification of the large fluxes of organic
N and NH4 –N by headwater rivers and streams
during episodic storm events that are largely
unrecorded.
The improved treatment of N cycling in conceptual models will lead to the improved treatment of
N cycling in mathematical models and massbalance studies. In turn, these improved scientific
data and models will help answer important
resource management and policy questions, such
as those posed in the Introduction.
The Illinois State Water Survey has developed
an N-cycle site which can be accessed at http:yy
www.sws.uiuc.eduynitroysplash.asp. Comments
and suggestions for improving the site are invited.
More detailed studies of N cycling in the ERW
are being conducted.
Acknowledgments
We thank reviewers who provided useful comments on earlier versions of the article, Linda
21
Hascall for the graphics and Eva Kingston for
editing. This work was supported by the Illinois
State Water Survey and by Illinois Council on
Food and Agriculture (C-FAR) grants 99Si-86-5A
and 02E-14-5.
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