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IV. Impact of Agricultural Practices on Subsurface Microbial Ecology

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face microorganisms carried out ecologically relevant processes (Section II.B.3).

This is because the water in aquifers, though considered subject to geochemical

evolution by geologists (Freeze and Cherry, 1979), has also been viewed as a

virtually unchanging resource. Prior to the era of manufacture and improper disposal of toxic contaminant compounds, water that made the journey from the atmosphere to the soil and infiltrated through to the water table (Fig. 2) was almost

unquestionably of drinking water quality. When subsequently drawn back to the

surface through wells, this water was still of drinking quality. With no evidence

(or need) for microbiologically mediated changes in the composition of water

while it resided in aquifers, there was no reason to propose that microorganisms

were present and metabolically active in the subsurface.

Why then have the subsurface habitat and subsurface microbial ecology been

recently explored? As mentioned in Section I, the major motivation for advancing

the science of subsurface microbiology has been to alleviate groundwater pollution. The key hope has been, and continues to be, that the potential of subsurface

microbial populations for eliminating toxic organic compounds will match that of

microorganisms in more thoroughly studied habitats such as surface water, sediments, and soils.







Perhaps the most efficient way to assess impacts of agricultural practices on

subsurface microbial ecology is to look for changes in the normal composition

and function of the native microflora. But “to look for change” requires two

pieces of prior knowledge. First, one needs to have sufficient familiarity of the

natural history of subsurface systems to know what “normal composition and

function” are. Unfortunately, the complexity of microbial communities in virtually all natural habitats has defied complete understanding by microbial ecologists

(Hobbie, 1993; Hobbie and Ford, 1993; Madsen, 1995). This general lack of understanding in microbial ecology is compounded in subsurface environments by

this habitat’s physical inaccessibility and the relatively brief period of time since

the inception of the modern era of subsurface microbial ecology in the 1970s

(Ghiorse and Wilson, 1988; Chapelle, 1993). There simply is no robust data base

on subsurface microbial communities that defines a “normal composition” of subsurface microflora. Moreover, the “normal function” of microorganisms in subsurface environments is uncertain because (as discussed earlier in Sections II.B.2,

II.B.3, and 1V.A) the biogeochemical processes acting in groundwater between

infiltration and discharge events have not traditionally been obvious. Thus, present

knowledge of the natural history of subsurface habitats does not let us definitively

state what “normal” structure and function are.



The second piece of prior knowledge required for assessing the impact of agricultural practices on subsurface microbial ecology is possession of reliable tools

to “look for changes.” In Sections 1I.B.I and 2 and Table I1 a hierarchy and potential responses of microorganisms (bacteria) to environmental changes were

outlined. The rich and varied mechanisms by which bacteria adapt to external

stimuli, especially at the gene, enzyme, and cellular levels of the hierarchy, are

remarkable. But equally remarkable is the uncertain applicability of these molecular response mechanisms to ecologically relevant, mixed microbial communities

such as those in groundwater. Most of the information presented in Table I1 and

Section 1I.B.I was derived from pure cultures, unlikely to be present in subsurface

microbial communities, that were grown under laboratory conditions that depart

radically from those found in field sites. Thus, the usefulness of information from

a few domesticated,pampered bacteria in interpreting the status of real-world ones

is highly suspect. This discontinuity between “pure” and “applied” research is

a common problem. For instance, Mayer et al. (1992), in discussing the use of

biomarkers as indicators of anthropogenic stresses, have emphasized that the

utility of biochemical, physiological, and histological measures in the field is unknown because the pertinent data base is derived from single organisms raised

under laboratory conditions. Mayer et al. ( 1992) emphasized that testing and

evaluation of biomarker methods in field settings are of critical importance.

Although the gene, enzyme, and cellular level measures described in Table I1

and Section 1I.B. 1 are of uncertain applicability to subsurface microbial communities, the population and community level measures have a long history of usage

in soil microbiology. Hicks et al. (1990) have reviewed both conceptual and practical aspects of attempts to evaluate the responses of microorganisms in soil to

environmental change (specifically to exposure to xenobiotic chemicals). As mentioned earlier in this section, the complexity of naturally occurring microbial communities is overwhelming and soil microbial communities are perhaps the best

example of this rule. The fundamental challenge in attempting to devise useful

procedures for testing for ecological effects of environmental change is selecting

meaningful parameters to measure. Naturally occurring microbial communities

are so diverse, resilient, and dynamic that individual tests are often woefully inadequate because they characterize the responses of only a small proportion of the

total microbial community present in any given environmental sample. The

strategy advocated by Hicks et al. (1990) maximizes the relevance of a testing

procedure by including a multiplicity of measures performed on soils in the laboratory and in the field. To achieve the desired breadth of tests, one should focus

on processes of major ecological relevance including carbon cycling via respiration; nitrogen cycling via nitrogen fixation, denitrification, nitrification, and ammonification; l4CO2 production from radiolabeled amendments of pesticides,

plant residues, or readily metabolizable carbon compounds such as glucose and

acetate; extracellular enzyme assays such as phosphatase, dehydrogenase, and



peroxidase; enumeration of common microbiological groups (i.e., bacteria, fungi,

actomycetes, protozoa, algae); and measures of overall metabolic status (ATP concentration, adenylate energy charge, etc.). Hicks et al. (1990) closed their review

article with several conclusions. Notable among these were:

1. “Agricultural chemicals do not appear to have any long-term harmful effects

on soil microbial activity when applied at recommended field levels.”

2. “When exposed to xenobiotic compounds, various segments of the soil microbial community are affected to different extents. The degree to which a xenobiotic affects microbial activities is largely dependent upon the chemical, its dosage and method of application, and the particular physicochemical characteristics

of the soil.”

3. “The most pressing issue is deciding which of the available (measurement)

methods are the most valid for investigating xenobiotic effects and how can they

be applied for best results.”

4. “A comprehensive system for evaluating the effects of xenobiotics on soil

microbial activity should be established.”

These four conclusions reflect the general condition of “effects testing” in microbial ecology. Regarding the specific objectives of this chapter, one can confidently modify these four conclusions by substituting “agricultural practices” for

both “agricultural chemicals” and “xenobiotics” and by substituting “subsurface microorganisms” for “soil microbial activity.” In other words, methods for

measuring responses of subsurface microorganisms to environmental change are





1. Agricultural Practices That Cause Physical Changes

in the Subsurface Habitats

Table IV provides a framework for understanding the types of agriculturally

induced changes in the subsurface habitat and their potential and documented impacts on subsurface microorganisms. There are essentially two different categories

of changes that can be imposed on the subsurface by agricultural practices: physical and geochemical. Physical changes are direct effects of water infiltration (as

discussed in Section I11 and Fig. 6, water infiltration is the primary means by

which agriculture influences the subsurface habitat), independent of solutes that

may be present. Geochemical changes result from solutes deposited in the subsurface after being entrained in water infiltrating through from the surface.

A widespread agricultural practice, surface irrigation, can readily increase the

flux of water infiltrating through to the subsurface. When Brockman et al. (1992)



Table IV

Types of Agriculturally Induced Changes in the Subsurface Habitat and Their Potential and

Documented Impacts on Subsurface Microorganisms

Type of change

Physical changes

Increased volume of water infiltrating to the subsurface

Decrease volume of water infiltrating to the subsurface

Alteration of soil structure to

increase or diminish



Geochemical changes

Increased ionic strength


Potential or documented

impact on subsurface


Shift in microbial community away from desiccation-resistant populations

Transport of microorganisms from soil into and

within the subsurface


Shift in microbial community toward desiccationresistant populations

Diminished transport of microorganisms from soil

into and within the subsurface habitat

Increased or diminished flux

of water infiltrating from

the soil may alter microbial populations (as


Warm surface waters infiltrating to shallow aquifers

may stimulate microbial

activity strictly because

of increased


Deposition of salts from irrigation waters may select

for salt-tolerant


Solutes present in infiltrating waters may affect

acid/base equilibria

which, in turn, may

stimulate or inhibit various populations in the microbial community

Selected references

Brockman er a/. (1992)

Ijzerman et d.( 1993); Harvey

and Garbedian (1991 ); Powelson er a/. (1990); Scholl

and Harvey ( 1992); Tim and

Mostaghimi (1991); Wan ri

a/. ( 1994)

Kieft et N / . ( 1993)

Ijzerman P I ol. ( I 993); Harvey

and Garbedian (1991); Powelson er a/. ( 1990); Scholl

and Harvey ( 1 992); Tim and

Mostaghimi ( I 99 I); Wan et

a/. (1994)







Table IV-Conrinurtl


Type of change

Specific solutes

Nitrate introduction

Pesticide introduction

Introduction o f other organic


Introduction of other nutrients


Potential or documented

impact on subsurface


Nitrate leaching into the

subsurface may satisfy

nutritional nitrogen requirements and/or act as

final electron acceptor

under anaerobic


Pesticides may serve as carbon and/or nitrogen

sources, act as toxins, or

have no impact on subsurface microorganisms

Crop root exudates and

compounds released from

decay of crop residues

may serve as carbon and

energy sources

Fertilizer components (especially phosphorus, perhaps potassium and urea)

may relieve nutrient limitations, hence stimulate

microbial growth and


Selected references

Smith er ul. (1991a,h); Starr

and Gillham (1993); Wilson

Pt a / . ( 1990)

Agertved (1992); Dippel et cil.

(1991); Konopka (1993);

Konopka and Turco ( 199 I );

McMahon et a/. ( 1992)



(Table IV) compared the microflora of two cliff face-derived sediments that were

geologically similar but had different moisture regimes, the viable microbial communities were also quite different. In these Columbia River Basin sediments, the

“control” field sediment (which received no irrigation water) was rich in actinomycetes (Gram-positive spore-forming bacteria). However, the field sample of

sediment that had received irrigation water was dominated by Gram-negative

nonspore-forming bacteria. Thus, infiltration of water at this arid field site appears

to have produced a substantive change in the composition of the subsurface microbial community. It is not certain precisely how the change in subsurface community was achieved. In all probability the community composition was altered

both because Gram-negative bacteria were transported into the sediment with the

water and because the altered moisture regime may have favored growth and survival of the introduced and indigenous Gram-negative bacteria.



The potential for microorganisms (especially bacteria and viruses) to be transported from surface to subsurface environments in both saturated and unsaturated

systems has been reasonably well documented (Table IV). The quantity of microorganisms entrained and transported is considered to be roughly proportional to

the flux of water.

As discussed in Section 111 and Table 111, the converse of irrigation is also a

routinely implemented agricultural practice. Under certain conditions it is conceivable that direction of moisture away from agricultural soils could cause a shift

in subsurface microbial populations toward those that tolerate desiccation. Although Kieft et al. (1993) (Table IV) did not work beneath agriculturally active

soil (instead the study area was approximately 400 m beneath the Nevada test

site), these researchers showed that the ambient low-moisture conditions found at

the study area correlated well with the physiological capacity of sample-derived

bacteria to tolerate desiccation.

The final two entries of the Physical Changes section of Table IV are purely

speculative. A variety of soil tillage and crop growth practices, especially passage

of farm equipment over the soil surface and deep penetration of plant roots, routinely alter soil structure, hence permeability characteristics (Section 111). Such

alterations in permeability can undoubtedly influence the volume of water reaching the subsurface and alter microbial populations either toward or away from

desiccation resistance as already discussed. Depending on climatic conditions,

heat capacities of subsurface sediments, and infiltration rates, it is also conceivable

that the warmth of rapidly infiltrating irrigation waters could stimulate microbial

activity in cooler subsurface zones. Such a temperature effect has never been

documented, however.

2. Agricultural Practices That Cause Geochemical Changes

in the Subsurface Habitat

In the Geochemical Change section of Table IV, increased ionic strength is a

type of environmental perturbation that is certain to have occurred throughout the

world in arid field sites under irrigation. Soil salinization represents one of the

major problems implicit in many modern agricultural practices (Frenkel and

Meiri, 1985; Magaritz and Nadler, 1993; McTernan and Mize, 1992; Umali,

1993). Despite the near certainty of salt deposition in arid climate subsurface environments, the impact of increased salinity on subsurface microorganisms has

not been investigated. However, at least one study designed to assess the impact

of salt stress on soil microorganisms has been completed. Kilham (1985) tested

a variety of stresses, only one of which (NaCI) was relevant to salinization processes. The investigator amended soil samples with "C-labeled glucose and measured both evolved W O z and extracted ['T]glucose after exposure to 0.3 M

NaCI. An additional metabolic measure was dehydrogenase activity. Kilham



( 1985) found that the NaCl amendment decreased both glucose respiration and

dehydrogenase activity. These same effects are likely to occur for subsurface


Much as was the case for increased ionic-strength effects on subsurface microorganisms above, it is likely that other solutes entrained in infiltration waters may

also alter the geochemical composition of groundwater. These compositional alterations may stem from mechanisms such as acidbase reactions or via contribution of both crop-derived and fertilizer-derived organic and inorganic nutrients.

However, neither of these geochemical changes nor their effects on subsurface

microbial communities are well documented (Table IV).

Nitrate contamination of groundwater is one of the most widespread environmental problems associated with agricultural practices (Spalding and Exner,

1993). Agricultural sources of nitrate include fertilizer application and animal manures. The major physiological impact of enriched concentrations of nitrate for

subsurface microorganisms is its utilization under anaerobic conditions as a final

electron acceptor. When used as a final electron acceptor, nitrate may undergo

microbial denitrification reactions to N2 or may be completely reduced to ammonia (Tiedje, 1988; 1994). Although the denitrification potential for subsurface

microorganisms is well documented (see Table I), only relatively rarely has in situ

denitrification activity by groundwater microorganisms been demonstrated (see

Korom, 1992). A selection of representative subsurface denitrification studies is

cited in Table IV. These studies are discussed next.

At a field site contaminated with sewage treatment effluent, Smith et al. (1991 b)

documented in situ denitrification activity by subsurface microorganisms by determining trace amounts of N,O (a key denitrification intermediary metabolite) in

water chemistry profiles from the site. An innovative study by Starr and Gillham

( 1993) examined denitrification by subsurface microorganisms in two agricultural

areas in Ontario. The acetylene block technique (which allows denitrification activity to be measured via accumulation of relatively high concentrations of N 2 0 )

was applied to cores of subsurface sediments incubated in the laboratory and to

sealed volumes of well water incubated in the field. In unamended in situ assays

at one of the study sites, groundwater microorganisms converted NO,- to N 2 0 ,

providing convincing evidence that the agriculturally derived NO, - underwent

microbially mediated denitrification reactions in situ. Starr and Gillham ( 1993)

also found that denitrification rates were sometimes limited by organic carbon

(which serves as an electron donor for the reaction). In addition, the authors concluded that the active subsurface microbial community was responsible for preventing migration of nitrate by removing it as it entered the site.

Additional evidence for in situ denitrification activity in groundwater beneath

fertilized agricultural soils has been obtained by Wilson et al. (1990). These researchers used measurements of trace gases and stable nitrogen isotopes to compute N2 to Ar ratios, microbial fractionation of I5N versus I4N, and historical re-



charge temperatures for groundwater samples along a landscape gradient. Wilson

et al. (1990) found that N2 to Ar ratios increased markedly down gradient; this

and related data allowed these investigators to infer that microorganisms had

reduced as much as 33 mg/liter of nitrate in the anaerobic zone of a limestone


The microbiological fate of synthetic pesticides in groundwater systems has

also been investigated to a limited extent (Table IV). Both Konopka and Turco

( 1 99 1) and McMahon et al. (1 992) conducted laboratory studies which incubated

subsurface sediments in the presence of radiolabeled herbicides (atrazine and/or

metolachlor). Despite the considerable duration of the experiments (up to

128 days) and the presence of [ 14C]glucosemineralization activity, neither Konopka and Turco ( I99 1 ) nor McMahon et al. (1992) were able to demonstrate microbially mediated conversion of the herbicides to COz. Using an innovative and

challenging field approach, Agertved et al. (1992) examined the behavior of two

herbicides, MCPP and atrazine, injected into groundwater beneath the Canadian

Forces Base in Borden, Ontario. The fate of the herbicides was monitored using a

conservative tracer (chloride), multilevel piezometers, and a partially enclosed

chamber installed in the field. Furthermore, laboratory incubations of sediment-groundwater mixtures amended with the herbicides were also prepared. After 74 days of incubation in the laboratory microcosms, Agertved et al. (1992)

failed to discern any herbicide biodegradation activity. Similarly, using a decline

in the ratio of atrazine to chloride in field tests to indicate microbial metabolism,

no subsurface biodegradation activity for atrazine was detected. However, Agertved et al. ( 1992) did report a decrease in the field concentration of MCPP relative

to chloride. Despite lack of corroboration by other tests, the authors attributed the

MCPP decline to in situ metabolism by subsurface microorganisms.

3. Testing Integrated Effects of Agricultural Practices

on SubsurfaceMicroorganisms

It is evident from these series of reports (Table IV) that nitrogen transformations, particularly denitrification reactions, have been thoroughly studied in

nitrate-enriched groundwaters beneath agricultural lands. Furthermore, these

reports indicate that pesticide biodegradation activity in groundwater has been

sought but, in general, not found. There is an analogy here between subsurface

denitrification/pesticide research and broader pollution abatement research already discussed in Sections II.A.2, II.B.2, and 1V.A. For denitrification, pesticide

metabolism and microbial detoxification processes as a whole, the concern motivating scientific investigations has not been “preservation of known and beneficial

subsurface microbial ecosystem processes” but rather “now that the contaminants

have escaped, are the indigenous subsurface microorganisms capable of detoxifying their habitat?” Implicit in these types of research approaches is a very narrow focus for discerning the impact of agricultural processes on subsurface micro-



bial ecology. Often the only aspect of subsurface microbial ecology that is truly

of interest is a single process (e.g., denitrification activity or biodegradation activity) that will undo the single environmental insult of interest (e.g., groundwater

contamination by nitrate, pesticides, or other organic compounds).

Section 1V.B and portions of Sections 1I.B.I and 2 reviewed the spectrum of

measures that have (somewhat unsatisfyingly) been developed and applied to soil

systems for evaluating impacts of environmental perturbations on microbial communities. As may be evident from the discussion in Sections 1V.C. 1 and 2, and the

paucity of listings in Table IV, such broad-based biochemical impact studies have

not been conducted on groundwater microbial communities. However, Van Beelen et al. ( 199 I ) have conducted limited tests on the effects of environmental pollutants on subsurface microflora. In this report, the field site from which the subsurface samples were derived was not thoroughly described, but uncontaminated

soils from depths up to 4 m were examined. Van Beelen et al. (199 1) used aerobic

and anaerobic respiration of I4C-labeled acetate and glucose to assess the response

of subsurface microorganisms to four model pollutant materials: acid (hydrogen

chloride), cadmium, chlorite, and a wood preservative (pentachlorophenol). Although inhibitory concentrations of these “environmental stressors” were measured, they had little, if any, relevance to agriculture practices.

In light of the dearth of data directly addressing the response of subsurface

microbial communities to agricultural practices, mention of related studies using

soil is deemed appropriate. Soil microbial communities are generally more robust

and diverse than subsurface communities (see Section II.A.2); thus, the results of

studies described next should be viewed as insensitive approximations of the behavior of subsurface microbial communities.

Wardle and Parkinson (1991) applied the herbicides 2,4-D and glyphosate to

soils in field plots; removed soil samples 1, 5, 15, and 45 days later; and at each

sampling time conducted laboratory-based respiration assays as well as enumerations of bacteria and fungi. After an extensive statistical analysis, Wardle and

Parkinson ( 1 991) concluded that glyphosate had no effect on soil microorganisms

and that the applied 2,4-D stimulated respiration for only the first 5 days. Beyond

this time period, 2,4-D had no influence.

In another study designed to evaluate the side effects of the herbicide Dinoseb

on soil microorganisms, Malkomes and Wohler (1983) applied Dinoseb to two

different soils (a loamy sand and a sandy loam) in the laboratory and in the field.

After a 1-month exposure period these investigators measured respiratory activity,

dehydrogenase activity, total ATP, nitrification, total mineral nitrogen, and decomposition of added straw. Malkomes and Wohler (1 983) found that the herbicide

had no measurable effect on the sandy loam soil. However, in the loamy sand, all

activities except straw decomposition were somewhat inhibited.

Smith et af. ( I99 la) provided an impressive and extensive review of long-term

(33-35 year) effects of herbicide application (2,4-D and MCPA) to agricultural

soils. Microbiological properties (including total biomass; respiratory activity; ni-



trogen transformation; urease and dehydrogenase activity; numbers of bacteria,

fungi, and actinomycetes; and pesticide metabolism assays) were measured for

soils that did and did not receive herbicide applications in the field. Smith et al.

( 1991a) concluded: ( I ) that minor short-term soil biochemical effects were seen

(slight inhibition of soil enzyme activities, bacteria, actinomycetes, and slight

stimulation of soil respiration by 2,4-D); but (2) that long-term effects of herbicide

usage had not affected soil microbiological populations to a significant extent; and

(3) that the exposed soil microorganisms had become metabolically adapted to the

herbicides, thus showing enhanced biodegradation activity.


Throughout this chapter an attempt has been made to elucidate relationships

between seemingly disparate disciplines that fall within a continuum from ecosystem ecology, geology, hydrology, agriculture, nutrient cycling, ecological effects

testing, and the ecology, physiology, and molecular biology of microorganisms.

An additional objective has been to use the relationships between these disciplines

to weave a coherent fabric addressing the impacts of agricultural practices on

subsurface microbial ecology. An itemized list describing the status of current

knowledge pertinent to impacts of agricultural practices on subsurface microbial

ecology is described next. These listed items are supported in sections of this

chapter noted in parentheses.

1. The methods of environmental microbiology and microbial ecology provide

only a “primitive ability” (Hobbie, 1993) for describing (a) the organisms present

in their microhabitats in nature; (b) what these organisms are actually doing; and

(c) what controls their activity and growth (Section 1V.B).

2. Given the relatively recent interest in exploring subsurface microbial

ecology (aseptic sampling procedures were developed only in the late 1970s), the

admittedly limited methodologies (item No. 1) have not been applied for a sufficient period to provide a robust data base for understanding the “normal” composition and function of subsurface microbial communities (Sections I, II.B.3, and


3. The lack of understanding in subsurface microbial ecology that follows from

item No. 2 poses obstacles for assessing the impact of any and all environmental

perturbations of subsurface microbial ecology, including agricultural practices

(Section 1V.B).

4. Despite limitations of the information describing the inhabitants of subsurface environments and both their potential and actual metabolic activity (item

Nos. l-3), new data and new types of data are arriving at an accelerating rate.

Thus our level of understanding is rapidly improving (Sections II.A.2, II.B.2 and

3, and 1V.C).



5. Ongoing developments in disciplines such as molecular biology and biochemistry (describing cellular- and subcellular-level processes) and ecosystemeffects testing (describing population- and community-level processes) offer much

promise for investigating the status of subsurface microbial communities in field

sites. Application of the new molecular biology and biochemical methods (derived from laboratory-grown single organisms) to real-world microbial communities needs to be aggressively pursued. When perfected these methodologies will

provide far-reaching insights (Sections 1I.B. 1 and 1V.B).

6. One simple reason for limited knowledge of the impacts of agricultural practices on subsurface microbial ecology is that very few scientific investigations that

directly address these issues have been conducted. This interdisciplinary research

area has not traditionally been a high priority in governmental research programs,

although this policy is probably changing both in North America and Europe

(Section 1V.C).

7. Regardless of the limitations of current knowledge describing how agricultural practices affect subsurface microbial ecology, decent progress in this area

has already been made. We know that infiltrating water is the primary means by

which materials are transported from soil into the subsurface. Furthermore, the

water itself and many of the entrained materials (microbial cells, salts, organic or

inorganic nutrients, pesticides, etc.) affect the status of subsurface microorganisms. In light of promising developments in microbial ecological methods and

increasing interest in the interface between agriculture and groundwater resources,

rapid advancements in this important interdisciplinary area are imminent (Sections LA, II.A.l, 111, and 1V.C).


The author expresses his appreciation for provocative and insightful discussions with W. C.

Ghiorse. These discussions assisted in the development of many ideas presented in this chapter. Research support during preparation ofthis manuscript was provided by the Air Force Office of Scientific

Research (Grants AFOSR-9 1-0436 and 93-NL-073) and the Cornell Biotechnology Program, which

is sponsored by the New York State Science and Technology Foundation (Grant NYSCAT 92054).

and consortium of industries and the National Science Foundation. 1 am grateful to Patti Lisk and

Shirley Cramer for expert manuscript preparation.


Acton. D. W., and Barker. J. F. 1992. In situ biodegradation potential of aromatic hydrocarbons in

anaerobic groundwaters. J. Contum. Hydro/. 9,325-352.

Agertved, J., Rugge. K., and Barker, J. F. 1992. Transformation of the herbicides MCPP and atrazine

under natural conditions. Ground Wut. 30,500-506.

Albrechtsen, H.-J., and Winding, A. 1992. Microbial biomass and activity in subsurface sediments

from Vejen, Denmark. Microh. Ecol. 23, 303-317.

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