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V. Agriculturc and Soil Health

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inefficiencies. Soils which are incapable of storing nutrients require excessive or

continuous addition of soluble nutrients for crop growth, to compensate for

losses and inefficiencies. Soluble nutrients in excess of plant and microbial needs

will pass beyond the reach of plant roots, with potential consequences to the

groundwater already discussed, compounded by the loss of valuable and possibly

dwindling nutrient resources.

Soil organisms must be acknowledged as key architects in nutrient turnover,

organic matter transformation, and physical engineering of soil structure (see

Fig. 1). The microbial populations of the soil alone encompass an enormous

diversity of bacteria, algae, fungi, protozoa, viruses, and actinomycetes. As

many as 10,000 different species may be found in a single gram of soil (Torsvik

et al., 1990), just a small sample of the nearly two million species of microorganisms thought to exist worldwide, with a range of form and function beyond

current capacity for comprehensive study. While the specific functions and interactions of the majority of these organisms are as yet poorly elucidated, their role

as functional groups in soil health regeneration and maintenance is becoming

increasingly clear (Kennedy and Papendick, 1995).

The microbial biomass is largely responsible for mineralization and turnover

of organic substrates (Killham, 1994). It includes both primary and secondary

decomposers, aerobic, anaerobic, and switch-hitting digestors, highly specialized consumers of gourmet delicacies and feeding trough generalists, finicky

occupants of outlandish environmental niches and highly adaptable opportunists,

hard-driven frenzied achievers and slow-metabolizing plodders, diners of rich,

fatty substrates and those eking out an existence gnawing on tough lignaceous

scrap. As a group, the community of microbial populations acts without regard

for the future, but instead responds quickly to favorable conditions, reproducing

and consuming with wild abandon until substrate limitations cause population

declines, victims of their collective gluttony. They are in turn cannibalized by

their surviving compatriots. The result is a continuous cyclic ballet of nutrient

uptake and release that enables less ephemeral life forms in the soil to be supplied

with their nutritional needs in a somewhat regulated way.

The role of larger soil organisms in maintenance of soil quality and health has

finally begun to receive much deserved attention in soil science circles with the

publication of several excellent review articles in recent years (Berry, 1994;

Linden et a l . , 1994; Stork and Eggleton, 1992). Soil fauna cover a range of soil

functions beyond that of the soil microbial community. Anderson ( 1988) classifies soil invertebrates into three categories, based primarily on size. Microfauna

are those less than 100 p m in diameter and include protozoa, nematodes, and

rotifers. They are the aquanauts of the soil, existing in water films around soil

particles and free water in soil pores. They function as secondary consumers,

feeding largely on bacteria and fungi, thereby speeding the turnover of microbial

biomass and their associated nutrients. The diversity in nematode function is



vast, spanning many different trophic levels, and nematode identification has

been suggested as an indicator of soil organism diversity and soil quality (Bohlen

and Edwards, 1994; Bongers, 1990; Neher et al., 1995; Parmelee and Alston,


The mesofauna, according to Anderson’s classification, are those invertebrates

100-200 p m in diameter and include mites, Collembola (springtails), and the

Enchytraeidae, or pot worms, as well as thousands of species of insects and

spiders. They tend to be omnivores, dining on microflora and fauna, as well as

other mesofauna and decomposing plant residues. In this way they speed organic

matter turnover directly, as well as indirectly, by fragmenting residues, thereby

increasing the surface area available for colonization by smaller organisms.

Enchytraeidae affect soil structure through creation of aggregates resulting from

fecal pellets and through burrowing activities.

Macrofauna are greater than 200 pm and include ants and termites as well as

the box-office stars of the underworld, the earthworms. Ants and termites can

have localized profound effects on soil structure, but earthworms are more ubiquitous and have become unwitting symbols of a healthy, living soil. They can

contribute in several ways to soil health. Most notably, earthworm burrows can

occupy as much as 1% of the soil volume (Kretzchmar, 1982), aiding in infiltration and flowthrough of water (Lee, 1985), as well as providing pathways for

root exploration and faunal habitat. Their feeding habits can help homogenize the

topsoil and, in the case of surface feeders, incorporate large amounts of surface

litter into deeper soil levels. Their digestive process releases nutrients and fragments of plant residues, leaving behind fertile casts and mucus burrow linings

(Berry, 1994).

The conditions favoring high earthworm populations overlap to a great degree

with those considered indicative of a healthy soil-good soil structure, adequate

moisture, sources of fresh organic material, and absence of certain pesticides.

There are several studies which in fact show them to be in considerably greater

abundance in natural ecosystems than in cultivated land (Barnes and Ellis, 1979;

Mackay and Kladivko, 1985) and higher under “sustainable” than conventional

management. Unfortunately, their use as an indicator of soil health is complicated by the fact that the conditions which cause them to be absent, or in low

numbers, do not correlate entirely with other indicators of soil health. For example, many will burrow into deep soil layers during cold or extended dry periods.

Although earthworms are usually present in highly productive soils, some highquality soils may be devoid of earthworm activity due to such factors as tillage or

environmental restrictions (Linden et al., 1994).

It is becoming increasingly obvious what the consequences of soil organic

matter loss are, that soil organisms are both the preservers and the destroyers of

soil organic matter, and that human intervention has a profound effect in orchestrating their activities. Clearly, a new vision of the fragile soil resource is needed.



The concept of the soil as a living organism, as discussed earlier, is not new

(Balfour, 1948). It is complementary to the Gaia Hypothesis articulated by James

Lovelock and Lynn Margulis in the 1960s (Lovelock, 1991) which envisions the

whole planet as a living creature, continually manipulating and adjusting existing

conditions to favor its own survival. The soil-as-organism model is useful for

conceptualizing the various functioning systems in the soil as analogous to animal respiratory, digestive, and circulatory systems. Perhaps a slightly more

appropriate paradigm is of the soil as a community. The difference in the community model is that it is largely self-contained-the outputs and waste products of

one group become the inputs and energy sources of another. There can then be

more complementarity of function than can occur at the single organism level.

Within stable communities, there is little loss of nutrients from the system and

outflows of water and energy are balanced by inflows, mostly from rainfall and

solar radiation. In this context the need for complex, diverse and overlapping

functional groups in the soil becomes apparent. There is a need for both generalists, which perform the bulk of everyday chores, and the specialists in the

community which fill specialized niches. In such a model, diversity itself may

serve as an indicator for soil health.

A follow-through of the soil-as-community paradigm is the idea of plant

nutrition being more efficient if cycled through a complex web of organisms and

their natural environment which is governed by rules that ensure the survival of

the whole community. This has been called the “feed the soil, not the crop” tenet

of sustainable or regenerative agriculture. Such plant nutrition seeks to mimic

natural ecosystems and relies on the yearly mineralization of organic materials by

soil microorganisms in response to fluctuating food sources, moisture, aeration,

and temperature. To function properly, it requires a continuing commitment to

adding sufficient and diverse organic residues, and to maintaining crop rotations

that maximize the presence of living roots throughout the year and synchronize

nutrient availability closely with crop needs. It demands a more complex management than current conventional agriculture, and necessitates a higher degree

of planning, but theoretically will lead to more efficient and environmentally

benign nutrient use.

A perhaps more profound outcome of a soil that functions as a living community would be the degree of resilience and stability that develops over time. The

dynamic combination of diverse-function populations, sufficient energy supplies,

and tight nutrient cycling would be expected to provide the basis in agricultural

soils for the kind of buffering capacity against environmental stresses seen in

equilibrium level natural ecosystems (Hillel, 1991; Soule and Piper, 1992). In a

system following this model, shortfalls in yearly nutrient inputs could be supplemented by stored nutrients in the organic matter or microbial populations. The

effect of disease and insect invasions would be minimized by a diverse group of

antagonistic organisms, and by the presence of a limited proportion of suscepti-



ble plants present at any given time. At a physical level, years of drought stress

would be ameliorated by higher water-holding capacity and more favorable conditions for root growth due to high organic matter, just as the impact of floods

would be lessened by good infiltration and drainage.

Swift (1994) proposed that assessments of production sustainability should be

based on two components-nondeclining crop yield trends, and stability of yield

from crop cycle to crop cycle. There is evidence that this sort of stability can in

fact be achieved in agricultural systems. After a 5-year transition period, a

comparison of conventionally grown crops and organically grown crops showed

that all systems had equivalent yields averaged over 9 years, but the organic

systems had less year-to-year variation (Hanson et a l . , 1990; Peters, 1994).

Dormaar et al. (1988) reported improved tolerance to drought stress in degraded

soils receiving animal manures relative to soils receiving only commercial fertilizer. In the Western Corn Belt of the U.S.A., Sahs and Lesoing (1985) found that

yields of rainfed corn (Zea mays L.) for organic management systems using

animal manures and/or crop rotations were higher than those for conventional

monoculture management with fertilizers and pesticides, especially during years

of high temperatures and water stress.



While terms for an agriculture that seeks to mimic natural ecosystems are

abundant, the term regenerative agriculture, coined by Robert Rodale, is perhaps

the most descriptive. Regenerative agricultural theory assumes that food production systems have caused some degradation of the natural resource base, and

seeks ways to restore or regenerate them toward their original state through

making maximum use of the internal resources available on farms (Rodale,

1984, 1995). The tenets of regenerative agriculture have never been explicitly

laid out. However, laying aside the social and economic aspects, in terms of

production systems alone, they are essentially the same as those associated with

sustainable or “biological” farming, namely:

Soil organic matter replenishment is the cornerstone to regenerating soil

health. Plant residues are left in the field or returned as compost as much as

possible. Animal production systems are designed to return manures to the

soil, either directly by pasturing, or by more efficient manure handling and

spreading systems. The necessary removal of organic material in the form of

harvested crops is compensated for by growing green manure crops or by

amending with compost, which may actually be composed of community food

waste, thus tightening the nutrient loop.

Living cover should be maintained throughout all or most of the year. This



provides plant roots which can take up soluble nutrients throughout the year,

further tightening nutrient cycles and decreasing loss. Living cover also protects

against erosion, provides habitat and substrate for soil organisms, and increases soil organic residue inputs. Although the feasibility of cover crops may

be limited in drier climates by the potential for competition for available water

with a grain crop, perennial soil cover is still an ideal to use as a guideline.

Diversity is critical at every level. Crops may be grown in polycultures, or

in alternating strips, or diversity may be achieved at the whole farm scale, with

complex rotations occumng in numerous small fields. Rotations are based on

progressions of plants with complementary water and nutrient needs, pest

susceptibilities, and root system types. This above-ground diversity may be

expected to harbor below-ground diversity in the soil microbial, and faunal

communities as there is a greater variety in food and nutrient sources available. Farm animals may also contribute to the diversity, fulfilling various

niches in nutrient cycling and waste disposal.

Inorganic fertilizers and pesticides should be reduced or eliminated. While

inorganic fertilizers may provide nutrients in similar or identical forms as

mineralized organic sources, they are discouraged because they have no direct

long-term soil enhancing properties. Certain plant nutrients need to be provided in inorganic form to restore losses from crop removal; in such cases

naturally occurring minerals are preferred because they can be applied in less

concentrated slow-release forms and commonly require less nonrenewable

energy for production and distribution. Pesticide reduction has a twofold

purpose-to protect farm employees from exposure to harmful substances,

and to avoid creating imbalances in communities of soil organisms.

Tillage should be minimized. Excessive tillage leads to increases in organic

matter decomposition due to physical disruption of aggregates, increased aeration and warming. While some form of soil disturbance may be required to

control weeds, less disruptive cultivation implements are favored and multiple

strategies for dealing with weed pressure are employed.

The theory and the practice of regenerative agriculture are rarely, if ever,

entirely meshed, but there are some signs that movement toward these ideals

does in fact lead to improvements in soil health. A comparison of organically and

conventionally managed tomato agroecosystems in California (Drinkwater et a l . ,

1995) showed that soils managed organically for at least 4 years had slightly

greater percentages of soil organic matter, lower soluble N concentrations, and

higher levels of microbial activity and potentially mineralizable N. A similar

study in New Zealand on paired biodynamically managed and conventional

farms found higher levels of microbial activity, soil organic matter contents, and

soil nutrient supplying capabilities on the biodynamic farms (Reganold et al.,

1993). When compared to a continuous grain system, an 8-year agroecological



rotation in Alberta, Canada, showed evidence of increases in total C, N , and P,

available N , P and K, CEC, microbial biomass, and microbial respiration (Wani

et al., 1994). The legume-based cropping system in the Rodale farming system

trial now exhibits higher organic matter content and microbial biomass (Wander

et al., 1994), greater water stable aggregates (Friedman, 1993), and reduced

nitrate leaching (Harris et al., 1994) as compared to the conventional system,

while maintaining equivalent yields.

Other authors have reported improvements in soil characteristics following

transition to more complex rotations including legumes (Angers and Mehuys,

1988; Doran and Smith, 1991; Doran and Werner, 1990; Kay et al., 1988), from

reducing tillage (Doran and Linn, 1994; Karlen et al., 1989; Angers et al.,

1992), or adding organic soil amendments (Dormaar et al., 1988).



Current agricultural practices provide an abundant and generally safe supply of

food and fiber at an inexpensive cost to the consumer. The cost of agricultural

products at the market, however, does not reflect the full cost of the agricultural

system. Environmental costs relating to deleterious consequences of contemporary agriculture, such as soil erosion, polluted water supplies, and poisoned

wildlife, are currently ignored under conventional agricultural accounting methods. Though these negative consequences may be more an attribute of a larger

economic model affecting agriculture than of agriculture itself, the environmental costs are nevertheless transferred from farmers to people in other places or

future time periods (Domanico et al., 1986).

Estimated environmental costs of agricultural production are significant. Annual off-site damage from soil erosion by water in the United States has been

estimated at over $7 billion (Pimentel et al., 1995; Ribaudo, 1989). Damage

includes costs associated with the loss of water’s value for recreation, decreased

water storage capacity, flooding, dredging ports and navigable rivers, and treating water for industrial and household use. Of the total soil erosion caused by

water in the United States, as much as 75% has been attributed to agricultural

sources (Pimentel et al., 1976). Wind erosion damage is generally considered to

be less severe than that by water, but may be substantial in arid regions. Damage

by wind erosion to households and businesses in New Mexico, where two-thirds

of the land is used for agriculture, has been estimated to range from $260 to $466

million annually (Piper and Huszar, 1989). Contamination of water by agrichemicals may be the most costly environmental consequence of agricultural

production. Annual damage by pesticides and fertilizers to water quality is suspected to range in the billions of dollars (Duda, 1985; Nielsen and Lee, 1987).

Costs associated with surface and ground water contamination from agrichemi-



cals include remediation and replacement of contaminated water, impairment of

human and animal health, and loss of water fauna and flora (CAST, 1992b;

National Research Council, 1993b).

Compared to off-site environmental damage, the value of changes in soil

health is much more difficult to quantify. This is primarily due to advances in

agricultural technology that have masked much of the yield-reducing impact of

soil degradation (Crosson, 1982, p. 184). Calculation of nutrient replacement

costs from erosion, however, shed some light on the economic magnitude of

agriculture’s impact on soil health. Using the approach of Willis and Evans

(1977), estimated loss of nitrogen, phosphorus, and potassium in soil eroded by

water would amount to over $6 billion annually in the United States. [This

calculation assumes an average soil nitrogen, phosphorus, and potassium content

of 0.15, 0.12, and 2.2921, respectively. Current fertilizer prices were used for

NH,NO,, P20,, and K 2 0 . Soil erosion by water estimated to be 6.9 metric tons

per hectare per year on 155 million hectares of cropland (Kellogg et af., 1994).]

These costs reflect just a portion of the economic burden that must be incurred by

farmers and consumers alike. The costs of soil organic matter loss and soil tilth

deterioration are also likely significant, but remain undefined (Bauer and Black,


Given contemporary agriculture’s estimated cost to the environment and soil

health, economic consideration of natural resources is clearly necessary to

achieve agricultural sustainability. This has motivated scientists to call for the

application of natural resource accounting methods to agricultural production

(Domanico et a / . , 1986; Tangley, 1986). This call has been addressed through

efforts by the World Resources Institute (Faeth rt ul., 1991), who have employed

natural resource accounting to incorporate factors of soil health, regional environmental impacts, farm profitability, and governmental policy to evaluate agricultural sustainability.

The method used by Faeth et ul. ( I99 I ) to quantify changes in soil health relies

upon interconnected ideas of sustainability, business income, and natural resource depreciation. Sustainability implies that economic activity should meet

current needs without foreclosing future options (WCED, 1987). Business income encompasses this notion of sustainability when defined as “the maximum

consumption in a certain period that does not reduce potential consumption in

future periods” (Edwards and Bell, 1961, after Faeth, 1993). By this standard,

then, agricultural accounting methods can only be accurate if depreciation in

natural resource assets (i.e., soil) is subtracted from net revenues along with the

more common forms of farming assets, like machinery and buildings. Faeth et

a / . (1991) followed this standard by calculating a soil depreciation allowance in

evaluating the economic performance of agricultural production systems. By

incorporating output from the Erosion-Productivity Impact Calculator (EPIC)

model, the allowance estimated future income losses over a 30-year period from



the impact of production on the soil resource as declines in crop yield (Williams

et al., 1989). Inclusion of the soil depreciation allowance in their evaluation of

economic performance resulted in a reduced net farm income of $62 per hectare

per year for Pennsylvania’s best conventional corn-soybean management. This

cost represents a significant loss of wealth in the natural resource base: a loss,

represented by degraded soil health, that is currently ignored by conventional

agricultural accounting methods.


Establishing an ongoing assessment of the condition and health of our soil

resources is vital to maintaining the sustainability of agriculture and civilization.

As discussed earlier, the failure of several earlier civilizations was sealed by their

disregard for the health of finite soil resources. In today’s energy- and

technology-intensive world, the need for maintaining the health of our soil resources is imperative to sustaining productivity for increasing populations and in

maintaining global function and balance. Assessment of soil quality and health is

invaluable in determining the sustainability of land management systems. A

framework for evaluation or an index of soil quality and health is needed to

identify problem production areas, to make realistic estimates of food production, to monitor changes in sustainability and environmental quality as related to

agricultural management, and to assist government agencies in formulating and

evaluating sustainable agricultural and other land-use policies (Acton, 1993;

Granatstein and Bezdicek, 1992). Effective identification of appropriate indicators for soil health assessment depends on the ability of any approach to consider

the multiple components of soil function, in particular, productivity and environmental well-being. Identification of indicators and assessment approaches is

further complicated by the multiplicity of physical, chemical, and biological

factors which control biogeochemical processes and their variation in intensity

over time and space (Larson and Pierce, 1991). Realistic assessment of soil

quality and health, however, requires consideration of the multiple functions of

soil and their relative importance as dictated by societal and ecological needs.

There is a great need both to determine the status of and to enhance our soil

resources. Assessment and monitoring of the quality and health of soils must also

provide opportunity to evaluate and redesign soil and land management systems

for sustainability. Standards of soil quality and health are needed to determine

what is sustainable and what is not, and to determine if soil management systems

are functioning at acceptable levels of performance. Recently, Doran and Parkin

(1994) identified nine research needs critical to assessment and enhancement of

soil quality. The two highest priority needs were: (i) Establishment of reference



guidelines and thresholds for indicators of soil quality that enable identification

of relationships between measured soil attributes and soil function which permit

valid comparisons across variations in climate, soils, landuse, and management

systems; and (ii) development of a practical index for on-site assessment of soil

quality and health for use by farmers, researchers, extension, and environmental

monitors that can also be used by resource managers and policy makers to

determine the sustainability of land management practices.


Assessing the health or quality of soil can be likened to a medical examination

for humans where certain measurements are taken as basic indicators of system

function (Larson and Pierce, 1991). In a medical exam, the physician takes

certain key measurements of body system function such as temperature, blood

pressure, pulse rate, and perhaps certain blood or urine chemistries. If these basic

health indicators are outside the commonly accepted ranges, more specific tests

can be conducted to help identify the cause of the problem and find a solution.

For example, excessively high blood pressure may indicate a potential for system

failure (death) through stroke or cardiac arrest. The problem of high blood

pressure may result from the lifestyle of the individual due to improper diet, lack

of exercise, or high stress level. To assess a dietary cause for high blood pressure, the physician may request a secondary blood chemistry test for cholesterol,

electrolytes, etc. Assessment of stress level as a causative factor for high blood

pressure is less straightforward and generally involves implementing some

change in lifestyle followed by periodic monitoring of blood pressure to assess

the effectiveness of the change. This is a good example of using a basic indicator

both to identify a problem and to monitor the effects of management on the health

of a system.

Applying this human health analogy to soil health is fairly straightforward.

Larson and Pierce (1991) proposed that a minimum data set (MDS) of soil

parameters be adopted for assessing the health of world soils, and that standardized methodologies and procedures be established to assess changes in the

quality of those factors. A set of basic indicators of soil quality and health has not

previously been defined, largely due to difficulty in defining soil quality and

health, the wide range over which soil indicators vary in magnitude and importance, and disagreement among scientists and soil and land managers over which

basic indicators should be measured. Acton and Padbury (1993) defined soil

quality attributes as measurable soil properties that influence the capacity of soil

to perform crop production or environmental functions. Soil attributes are useful

in defining soil quality criteria and serve as indicators of change in quality.

Attributes that are most sensitive to management are most desirable as indicators

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