&EPA
United States
Environmental Protection
Agency
Office of Water
Office of Science and Technology
Washington, DC 20460
EPA-822-R-08-022
December 2008
www.epa.gov
METHODS FOR EVALUATING WETLAND CONDITION
#18 Biogeochemical Indicators
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United States Office of Water EPA-822-R-08-022
Environmental Protection Office of Science and Technology December 2008
Agency Washington, DC 20460 www.epa.gov
METHODS FOR EVALUATING WETLAND CONDITION
#18 Biogeochemical Indicators
Major Contributors
University of Florida, Institute of Food and Agriculture,
Soil and Water Science Department
K. Ramesh Reddy and Mark W. Clark
Prepared jointly by:
The U.S. Environmental Protection Agency
Health and Ecological Criteria Division (Office of Science and Technology)
and
Wetlands Division (Office of Wetlands, Oceans, and Watersheds)
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NOTICE
The material in this document has been subjected to U.S. Environmental Protection
Agency (EPA) technical review and has been approved for publication as an EPA document.
The information contained herein is offered to the reader as a review of the "state of the
science" concerning wetland bioassessment and nutrient enrichment and is not intended to
be prescriptive guidance or firm advice. Mention of trade names, products or services does
not convey, and should not be interpreted as conveying official EPA approval, endorsement,
or recommendation.
APPROPRIATE CITATION
U.S. EPA. 2008.Methods'forEvaluating WetlandCondition:BiogeochemicalIndicators. Office
of Water, U.S. Environmental Protection Agency, Washington, DC. EPA-822-R-08-022.
ACKNOWLEDGEMENTS
EPA acknowledges the contributions of K. Ramesh Reddy and Mark W. Clark,
both of University of Florida for writing this module.
This entire document can be downloaded from the following U.S. EPA websites:
http://www.epa.gov/waterscience/criteria/wetlands/
http://www.epa.gov/owow/wetlands/bawwg/publicat.html
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CONTENTS
FOREWORD vi
LIST OF "METHODS FOR EVALUATING WETLAND
CONDITION" MODULES vn
SUMMARY 1
PURPOSE 1
INTRODUCTION 1
CHARACTERISTICS OF THE SOIL AND WATER COLUMN 4
BASIC ELEMENTAL CYCLES 5
REFERENCE WETLANDS 11
SAMPLING PROTOCOL 12
WATER SAMPLING 15
SOIL SAMPLING 15
BlOGEOCHEMICAL INDICATORS 17
WATER QUALITY INDICATORS 18
SOIL QUALITY INDICATORS 23
MINIMUM MONITORING REQUIREMENTS 32
CASE STUDY: THE EVERGLADES 33
REFERENCES 39
LIST OF FIGURES
FIGURE 1: SCHEMATIC SHOWING THE LINKAGE BETWEEN
BlOGEOCHEMICAL PROCESSES AND INDICATORS 2
FIGURE 2: SCHEMATIC SHOWING BASIC NUTRIENT CYCLES IN
SOIL-WATER COLUMN OF A WETLAND 4
FIGURE 3: SCHEMATIC SHOWING BASIC CARBON CYCLES IN
THE SOIL-WATER COLUMN OF A WETLAND 6
IV
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FIGURE 4: SCHEMATIC SHOWING BASIC NITROGEN CYCLES IN
THE SOIL-WATER COLUMN OF A WETLAND 7
FIGURE 5: SCHEMATIC SHOWING BASIC PHOSPHOROUS CYCLES IN
THE SOIL-WATER COLUMN OF A WETLAND 8
FIGURE 6: SCHEMATIC SHOWING BASIC SULFUR CYCLES IN
THE SOIL-WATER COLUMN OF A WETLAND 1O
FIGURE 7: SCHEMATIC SHOWING THE PROPOSED STEPS TO DEVELOP
AND EVALUATE BIOGEOCHEMICAL INDICATOR CRITERIA 12
FIGURE 8.: AN EXAMPLE OF THE SAMPLING SCHEME USED TO PARTIALLY
QUANTIFY WITHIN WETLAND BIOGEOCHEMICAL GRADIENTS
WHILE MINIMIZING WITHIN ZONE VARIABILITY AND NUMBER
OF SAMPLES FOR ANALYSIS 13
LIST OF TABLES
TABLE 1: POTENTIAL WATER AND SOIL QUALITY INDICATORS
FOR ASSESSING NUTRIENT IMPACTS IN WETLANDS.
CDENOTES MINIMUM DATA REQUIRED FOR EACH SITE) 16
TABLE 2.A'. BACKGROUND CONCENTRATIONS (STANDARD ERROR)
SELECTED BIOGEOCHEMICAL INDICATORS IN FLOC/DETRITAL
COMPONENT OF VARIOUS HYDROLOGIC UNITS OF THE EVERGLADES.
THE PMN AND PMP REFER TO POTENTIALLY MINERALIZABLE
N AND P, RESPECTIVELY (WRIGHT, ET AL. 2OO2) 34
TABLE 2s: BACKGROUND CONCENTRATIONS (STANDARD ERROR) SELECTED
BIOGEOCHEMICAL INDICATORS IN SURFACE SOIL (O— 3 CM)
COMPONENT OF VARIOUS HYDROLOGIC UNITS OF THE EVERGLADES.
THE PMN AND PMP REFER TO POTENTIALLY MINERALIZABLE
N AND P, RESPECTIVELY. (WRIGHT, ET AL. 2OO2) 35
TABLE 3: IMPACT INDICES AND RELATIVE SENSITIVITY OF VARIOUS
BIOGEOCHEMICAL PROCESSES/INDICATORS MEASURED IN
DETRITAL AND SOIL LAYERS AT IMPACTED AND REFERENCE
SITES IN THE EVERGLADES. D = DETRITAL LAYER AND
S = O-1O CM SOIL (REDDY, ETAL. 1999) 37
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FOREWORD
In 1999, the U.S. Environmental Protection Agency (EPA) began work on this series of
reports entitled Methods for Evaluating Wetland Condition. The purpose of these reports is
to help States and Tribes develop methods to evaluate (1) the overall ecological condition of
wetlands using biological assessments and (2) nutrient enrichment of wetlands, which is one
of the primary stressors damaging wetlands in many parts of the country. This information
is intended to serve as a starting point for States and Tribes to eventually establish biological
and nutrient water quality criteria specifically refined for wetland waterbodies.
This purpose was to be accomplished by providing a series of "state of the science" modules
concerning wetland bioassessment as well as the nutrient enrichment of wetlands. The
individual module format was used instead of one large publication to facilitate the addition
of other reports as wetland science progresses and wetlands are further incorporated into
water quality programs. Also, this modular approach allows EPA to revise reports without
having to reprint them all. A list of the inaugural set of 20 modules can be found at the end
of this section.
This last set of reports is the product of a collaborative effort between EPAs Health and
Ecological Criteria Division of the Office of Science and Technology (OST) and the Wetlands
Division of the Office of Wetlands, Oceans and Watersheds (OWOW). The reports were
initiated with the support and oversight of Thomas J. Danielson then of OWOW, Amanda
K. Parker and Susan K. Jackson (OST), and seen to completion by Ifeyinwa F. Davis (OST).
EPA relied on the input and expertise of the contributing authors to publish the remaining
modules.
More information about biological and nutrient criteria is available at the following
EPA website:
http://www.epa.gov/ost/standards
More information about wetland biological assessments is available at the following
EPA website:
http://www.epa.gov/owow/wetlands/bawwg
VI
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LIST OF "METHODS FOR EVALUATING WETLAND
CONDITION" MODULES
MODULE # MODULE TITLE
1 INTRODUCTION TO WETLAND BIOLOGICAL ASSESSMENT
2 INTRODUCTION TO WETLAND NUTRIENT ASSESSMENT
3 THE STATE OF WETLAND SCIENCE
4 STUDY DESIGN FOR MONITORING WETLANDS
5 ADMINISTRATIVE FRAMEWORK FOR THE IMPLEMENTATION OF
A WETLAND BIOASSESSMENT PROGRAM
6 DEVELOPING METRICS AND INDEXES OF BIOLOGICAL INTEGRITY
7 WETLANDS CLASSIFICATION
8 VOLUNTEERS AND WETLAND BIOMONITORING
9 DEVELOPING AN INVERTEBRATE INDEX OF BIOLOGICAL
INTEGRITY FOR WETLANDS
1O USING VEGETATION TO ASSESS ENVIRONMENTAL CONDITIONS
IN WETLANDS
11 USING ALGAE TO ASSESS ENVIRONMENTAL CONDITIONS IN WETLANDS
12 USING AMPHIBIANS IN BlOASSESSMENTS OF WETLANDS
13 BIOLOGICAL ASSESSMENT METHODS FOR BIRDS
14 WETLAND BIOASSESSMENT CASE STUDIES
15 BIOASSESSMENT METHODS FOR FISH
16 VEGETATION-BASED INDICATORS OF WETLAND NUTRIENT ENRICHMENT
17 LAND-USE CHARACTERIZATION FOR NUTRIENT AND SEDIMENT
RISK ASSESSMENT
18 BlOGEOCHEMICAL INDICATORS
19 NUTRIENT LOADING
2O WETLAND HYDROLOGY
Vll
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SUMMARY
PURPOSE
rhis module discusses biogeochemical
parameters and their use as indices to
characterize the nutrient status of wetlands.
Biogeochemical processes in the soil and wa-
ter column drive key ecosystem functions
associated with wetland values (e.g. water
quality improvement through denitrification,
long-term nutrient storage in the organic mat-
ter, etc.). Process level measurements reflect
the functionality of a wetland and potential
impairment due to impacts; however, these
measurements are often tedious and costly.
Instead, it is possible to develop a relative
measure of process rates and potential by
evaluating components of biogeochemical
cycles that are either end products or sources
of material for a given process. In the case
of many processes within the carbon (C),
nitrogen (N), phosphorus (P) or sulfur (S)
cycles, microbial communities mediate the
rate and extent of these reactions in soil and
the water column. As a result, biogeochemi-
cal indicators associated with these processes
often respond rapidly to perturbations and
are spatially restricted to the impact zone.
These indicators will persist over moderate
time scales and in the absence of standing
water. In the discussion that follows, we will
introduce a suite of basic and more advanced
biogeochemical indicators that can be used to
characterize wetland condition and evaluate
those indices. We will also examine indices
that have been most sensitive to nutrient im-
pact in the Florida Everglades.
rhe purpose of this module is to describe
biogeochemical indicators that can be
used to monitor nutrients in wetland systems
in an ecologically meaningful and scientifi-
cally rigorous manner.
INTRODUCTION
Many species depend upon wetlands for
successful completion of their life cy-
cle and most require, or benefit from, nearby
aquatic habitat. Changes in the structure and
function of a wetland will eventually have
far-reaching effects on the biota of the wet-
land and the surrounding uplands. Monitor-
ing wetland systems can provide information
on environmental change, including changes
in community structure and function, in both
the wetland and adjacent upland watershed.
Development of sound concepts and meth-
odologies for ecosystem monitoring require
an understanding of ecosystem structure and
function. Given the high cost of environ-
mental monitoring in terms of time, human
resources and funding, methods need to be de-
veloped that are simple, efficient, scientifically
rigorous and ecologically meaningful. One of
the most attractive approaches to developing
scientifically rigorous methods is based on
the concept of using physical, chemical or bi-
ological properties or processes as indicators
of wetland condition, change or response to
anthropogenic impacts.
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Wetlands host complex microbial commu-
nities, including bacteria, fungi, protozoa,
and viruses. The size and diversity of micro-
bial communities are directly related to the
quality and the quantity of resources avail-
able in the system. Many of the water and
soil parameters that influence the ecosystem
are the end products of biogeochemical pro-
cesses that are microbiologically mediated.
Microbial processes and populations often
have more rapid turnover times than higher
trophic stages. Due to their large size, they
are often more responsive to environmental
changes at lower thresholds. Microbial pro-
cesses are potentially very sensitive to per-
turbations such as external nutrient loading,
hydrologic alterations, and fire. These char-
acteristics make them efficient indicators of
wetland conditions.
Biogeochemical processes are also like-
ly to be highly reliable indices in the sense
that ecological changes at such a fundamen-
tal level will affect all species utilizing the
ecosystem. Changes at higher levels, such as
a decline in populations of a suite of higher
organisms may be due to factors that affect
only a small portion of the biota, whereas
changes in biogeochemical processes signify
comprehensive alteration of the biota. Thus,
when describing the structure and function
of an ecosystem, it is critical to evaluate the
water and soil quality using an integrated
framework that links processes and associ-
ated indicators. Biogeochemical processes
may be sensitive and reliable indicators of
wetland condition, but their measurement can
be time-consuming and expensive. However,
concentrations of certain chemical substrates,
intermediates, and end products of ecologi-
cally potent, biogeochemical processes may
provide rapid and inexpensive indicators of
Stressors at Varied Spatial
and Temporal Scales
FIGURE 1: SCHEMATIC SHOWING THE
LINKAGE BETWEEN BIOGEOCHEMICAL
PROCESSES AND INDICATORS.
the rates of those processes. Hence, simple
measurements of biogeochemical processes
in wetlands could be extremely efficient in-
dicators of the wetland and the entire water-
shed. Furthermore, relationships between in-
dicators and processes may provide a more
reliable estimate of ecosystem health for as-
sessment at a landscape level.
A fundamental understanding of the bio-
geochemical processes regulating the func-
tions of the ecosystem is critical to evaluating
nutrient impacts and successes of restoration
efforts. The certainty associated with an as-
sessment decreases if the factors that affect
biogeochemical processes regulating the fate
and transport of nutrients in wetlands are
not well understood; i.e., the risk assessment
is only as good as the information/knowl-
edge available at the time. Therefore, it is
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imperative that sound linkages are developed
between nutrient indicator parameters and
assessment of structural and functional im-
pacts.
Eutrophication of wetlands can be attrib-
uted to: (i) increased external inputs of nutri-
ents from point and non-point sources; and/
or, (ii) accelerated nutrient cycling within the
soil associated with change in environmental
conditions of the soil and water column. Eu-
trophication is often linked to only external
sources of nutrients in many wetlands. How-
ever, internal nutrient sources can be equally
important, especially in highly impacted wet-
lands or older wetlands with large reserves of
organic and inorganic bound nutrients (Fisher
and Reddy 2001).
Anthropogenic nutrient loading from point
or non-point sources to a nutrient-limited
wetland system can alter physical, chemical
and biological properties and processes in
the soil and water that in turn can influence
ecosystem function and productivity (Figure
1). Wetlands, as low-lying areas in the land-
scape, receive inputs from all hydrologically
connected uplands. Many wetlands are open
systems receiving inputs of carbon (C) and
nutrients from upstream portions of the wa-
tershed including agricultural and urban ar-
eas. Prolonged nutrient loading to wetlands
can result in distinct gradients in floodwater
and soil. Mass loading and hydraulic reten-
tion time determine the degree and extent of
nutrient enrichment. The continual nutrient
loading in an oligotrophic wetland results in
two distinct zones: (i) a zone of high nutrient
availability or non-limiting nutrient condi-
tions near the input; and, a (ii) a zone of low
nutrient availability or nutrient limiting con-
ditions furthest from the input point. Between
these two extremes, there exists a gradient in
quality and quantity of organic matter, nutri-
ent accumulation, microbial and macrobiotic
communities, composition, and biogeochemi-
cal cycles. This enrichment effect can be seen
in many freshwater wetlands, most notably in
the subtropical Everglades (Davis 1991, Red-
dy, et al.1993; Craft and Richardson 1993 a, b;
DeBusk, et al.1994).
Low-nutrient systems are characterized by
low external loading of nutrients and relative-
ly closed, efficient elemental cycling (Odum
1969, 1985). Nutrients are held in tight, closed
cycles, whose efficiency enables maintenance
of energy flow. Hence, microbial activity and
plant productivity are nutrient limited. In re-
sponse, wetland vascular plants, periphyton,
and microbial communities are extremely ef-
ficient in utilizing and conserving nutrients.
Plant detritus in low-nutrient wetlands gener-
ally has high C:N:P mass ratios. The overall
turnover rate of high C:N:P ratio organic mat-
ter is usually slow, and long-term decompo-
sition may be both carbon and nutrient lim-
ited (Davis 1991; DeBusk and Reddy 1998).
Decomposition of high C:N:P plant detrital
material results in conditions where micro-
bial and periphytic communities out-compete
vascular plants for nutrients. However, envi-
ronmental factors such as water-table fluctua-
tions and fire can result in pulsed release of
nutrients, which may provide a significant,
although temporally infrequent, source of
plant- available nutrients in low- nutrient sys-
tems (Lodge, et al.1994).
High-nutrient wetland systems are char-
acterized by rapid turnover of C and nu-
trients, and by open elemental cycling,
where nutrient inputs often exceed demand.
These systems are low nutrient stressed and
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contain varying degrees of internal cycling.
In response, vascular plants and microbial/
periphyton communities are less efficient in
nutrient utilization. Plant detritus in impact-
ed areas generally has low C:N:P mass ratios,
and high net mineralization. The release of
nutrients during decomposition results in de-
creased importance of internal cycling by mi-
crobes and plants as compared with nutrient
loading from external sources.
One of the most attractive approaches to
measure and quantify nutrient availability is
based on the concept of using physical, chemi-
cal or biological properties or processes in the
soil and water column as indicators of change
or response to anthropogenic impacts. In this
report, we describe easily measurable indica-
tors with reasonable scientific rigor and reli-
ability that will assess nutrient conditions in
wetlands, then suggest a subset of indicators
that are most responsive and sensitive to be
used as endpoints for assessment of impacts.
As defined, assessment endpoints are explic-
it expressions of an environmental value to
be protected, while measurement endpoints
are measurable responses of an assessment
endpoint to a stressor (USEPA 1992; Suter,
1990).
In this module, we describe two levels of
indicators;(i) Level I indicators which are
easily measurable; and (ii) Level II indicators
which provide more scientific rigor and are
used to support easily measurable indicators.
Only select Level I indicators will be pro-
moted as indicator endpoints for assessment
of impacts.
NUTRIENT CYCLE IN WETLANDS
$
s
^ *v Bioavailable
nutrients
Nutrient Cycling in
Soil-Water Column
FIGURE 2: SCHEMATIC SHOWING
BASIC NUTRIENT CYCLES IN SOIL-
WATER COLUMN OF A WETLAND.
CHARACTERISTICS OF THE
SOIL AND WATER COLUMN
Aerobic-anaerobic interfaces in the water
ylcolumn (e.g. surfaces of detrital plant tis-
sue and benthic periphyton mats), at the soil-
water interface, and in the root zone of aquatic
macrophytes are unique features of wetlands,
as compared to upland landscapes. We define
the soil-water column, as soil with overlying
floodwater. The juxtaposition of aerobic and
anaerobic zones in wetlands supports a wide
range of microbial populations and associ-
ated metabolic activities, with oxygen-reduc-
tion occurring in the aerobic interface of the
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substrate, and reduction of alternate electron
acceptors occurring in the anaerobic zone
(Reddy and D'Angelo 1994, 1996). Under
continuously saturated soil conditions, verti-
cal layering of different metabolic activities
can be present, with oxygen-reduction occur-
ring at and just below the soil-floodwater in-
terface. Much of the aerobic decomposition
of plant detritus occurs in the water column;
however, the supply of oxygen may be insuf-
ficient to meet demands and drive certain
microbial groups to utilize alternate electron
acceptors, e.g., nitrate, oxidized forms of Fe
and Mn, sulfate and HCO3.
Soil drainage adds oxygen to the soil,
while other inorganic electron acceptors
may be added through hydraulic loading to
the system. Draining wetland soil acceler-
ates organic matter decomposition due to
the introduction of oxygen deeper into the
profile. In many wetlands, the influence of
NO3", and oxidized forms of Mn and Fe on
organic matter decomposition is minimal, as
the concentration of these electron acceptors
is usually low. The demand for electron ac-
ceptors of greater reduction potentials (NO3~,
Fe and Mn) is high and they are depleted rap-
idly from the system. Long-term sustainable
microbial activity is then supported by elec-
tron acceptors of lower reduction potentials
(sulfate and HCO3). Methanogenesis is often
viewed as the terminal step in anaerobic de-
composition in freshwater wetlands, whereas
sulfate reduction is viewed as the dominant
process in coastal wetlands. However, both
processes can function simultaneously in the
same ecosystem and compete for available
substrates.
A simple way to characterize wetlands for
aerobic and anaerobic zones is to determine
the oxidation-reduction or redox potential
(Eh) of the soil-water column. The redox po-
tential is expressed in units of millivolts (mV)
and is measured using a voltmeter. Typically,
wetlands with Eh values >300 mV are con-
sidered aerobic and exhibit drained soil con-
ditions, while soils with Eh values <300 mV
are considered as anaerobic and are devoid of
molecular oxygen.
BASIC ELEMENTAL CYCLES
CARBON
Compared to upland systems, most wet-
land ecosystems show an accumulation of or-
ganic matter, and therefore wetlands function
as global sinks for carbon. Accumulation of
organic (C) in wetlands is primarily a result
of the balance of two processes—C fixation
through photosynthesis and C losses through
decomposition. Rates of photosynthesis in
wetlands are typically higher than other
ecosystems, and rates of decomposition are
typically lower due to anaerobic conditions,
hence organic matter tends to accumulate. In
addition to maintaining proper functioning of
wetlands, organic matter storage also plays an
important role in regulating other ecosystems
and the biosphere. For example, up to 55%
of organic matter contains variable amounts
of nutrients N, P, S, O, and H. Therefore, the
accumulation of organic matter in wetlands
prevents transport of these nutrients to down-
stream aquatic systems.
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Carbon Cycle in Wetlands
•j*—*-41-»_<-»-»—WU—j-'J-
Decomposition/leaching I
Decomposition
leaching
Decomposition/leaching
FIGURE 3: SCHEMATIC SHOWING BASIC CARBON
CYCLES IN THE SOIL-WATER COLUMN OF A WETLAND.
Major pools of C storage include soil or-
ganic matter, detrital organic matter, micro-
bial biomass, plant biomass, and dissolved
organic carbon. Microbial biomass accounts
for <5% of the total C in the soil and water
column. The dominant constituents of soil
organic matter include cellulose, hemicellu-
lose, and lignin. The cellulose and hemicel-
lulose are readily biodegradable, while lignin
is highly stable under anaerobic conditions.
Carbon compounds recalcitrant to aerobic
and anaerobic decomposition tend to accu-
mulate in wetlands as either undecomposed
plant tissues (peat) or humic substances. For-
mation of humic substances probably involve
condensation reactions between reactive
phenolic groups of tannins and lignins with
water soluble nonhumic substances, which is
catalyzed by phenoloxidase enzymes in soils
(Francois 1990). This mechanism accounts
for the large molecular weight and heteroge-
neous humic substances that contain signifi-
cant amounts of N, P, S in their structures.
In the absence of oxygen, humic substances
are resistant to decomposition, and represent
a significant carbon and nutrient storage in
wetlands. Under drained conditions, humic
substances are more readily degraded, which
releases nutrients to the bioavailable pool,
thereby affecting downstream water quality.
Although the loading of anthropogenic
nutrients stimulates the growth of aquatic
vegetation in wetlands, a significant por-
6
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Nitrogen Cycle in Wetlands
Inflow
, i ™ 2"vy
k Nitrogen Fixation
— *"*. _ *••» _ ~-«_ »'4<— ••••*•
•NN2,N 20(g)
FIGURE 4: SCHEMATIC SHOWING BASIC NITROGEN
CYCLES IN THE SOIL-WATER COLUMN OF A WETLAND.
tion of the nutrient requirements may be met
through remineralization during decompo-
sition of organic matter. The rate of organic
matter turnover and nutrient regeneration is
influenced by: (i) hydroperiod (Happell and
Chanton 1993); (ii) characteristics of organic
substrates, (Webster and Benfield 1986; En-
riquez, et al.1993; DeBusk and Reddy 1998);
(iii) supply of electron acceptors (D An-
gelo and Reddy 1994a, b); and, (iv) addition
of growth limiting nutrients (Button 1985;
McKinley and Vestal 1992; Amador and
Jones 1995). For this reason both internal and
external factors can affect carbon cycling and
nutrient release from organic matter in wet-
lands.
NITROGEN
Nitrogen (N) enters wetlands in organic
and inorganic forms. The relative proportion
of each depends on source and type of wa-
ter entering these systems. Organic forms are
present in dissolved and particulate fractions,
while inorganic N (NH4-N, NO3-N and NO2-
N) is present in dissolved fractions. Particu-
late fractions are removed through settling
and burial, while the removal of dissolved
forms is regulated by various biogeochemi-
cal reactions functioning in the soil and water
column. Relative rates of these processes are
affected by physico-chemical and biological
characteristics of soils, organic substrates,
and water column.
7
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Nitrogen reactions in wetlands effectively
process inorganic N through nitrification and
denitrification, ammonia volatilization and
plant uptake. These processes aid in lower-
ing levels of inorganic N in the water column.
A significant portion of dissolved organic N
assimilated by plants returns to the water col-
umn during breakdown of detrital tissue or
soil organic matter, and the majority of this
dissolved organic N is resistant to decompo-
sition. Under these conditions, water leaving
the wetlands may contain elevated levels of
N in organic form. However, relative rates of
these reactions will depend on the optimal
environmental conditions present in soil and
water column.
PHOSPHORUS
Wetlands regulate phosphorus (P) reten-
tion by physical mechanisms (sedimentation
and entrainment) and biological mechanisms
(uptake and release by vegetation, periphyton
and microorganisms). Phosphorus in the in-
fluent water is typically in the form of soluble
and particulate fractions, with both forms
containing a certain proportion of inorganic
and organic pools. The relative proportions
of these pools depend on the source and the
type of water that enters the system. For ex-
ample, municipal wastewater may contain
a large proportion (>75%) as inorganic P in
soluble forms, as compared to effluents from
agricultural watersheds where percentage of
P loading is predominantly in the particulate
fraction. Constructed and riparian (buffer)
wetlands can function as effective sediment
traps, as such, P associated with suspended
sediments can be effectively removed by wet-
lands.
Phosphorus Cycle in Wetlands
Inflow
Outflow
FIGURE 5: SCHEMATIC SHOWING BASIC PHOSPHOROUS
CYCLES IN THE SOIL-WATER COLUMN OF A WETLAND.
8
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Inorganic P — content of wetland soils
vary with soil type. Typically, mineral wet-
land soils have a higher proportion of total
soil P in inorganic fractions. Discrete frac-
tions of inorganic P are determined using
simple chemical fractionation schemes (Red-
dy, etal.!998a).
Porewater P — Phosphorus present in dis-
solved form in soil interstitial water. Phos-
phorus present in this pool is highly mobile
and readily available for plant uptake or flux
into the overlying water column. The concen-
tration of P in soil porewater is regulated by
soil physicochemical properties and the soil's
capacity to adsorb or desorb P.
Exchangeable P — Inorganic P ions ad-
sorbed on positively charged soil or organic
matter surfaces. Under acid conditions, ex-
cessive protons may result in positive charg-
es on soil particle surfaces. These positive-
ly charged surfaces can adsorb negatively
charged phosphate ions. Phosphate sorbed on
surfaces maintains equilibrium with phos-
phate ions in soil porewater.
Iron and Aluminum bound P — Iron and
aluminum minerals can retain P by adsorbing
on their surfaces or forming discrete miner-
als, such as ferric phosphate or aluminum
phosphate. These minerals are stable under
acid soil conditions.
Calcium and Magnesium bound P — In
alkaline soils, inorganic P is typically bound
to calcium or magnesium based minerals.
Residue P — This form of P is insoluble in
either alkali or in acid. Phosphorus present in
this pool is typically not bioavailable.
Organic P fraction in soils and sediments
accounts for a high proportion of the total
P, with this factor accounting for 20-60% of
total P in mineral soils (Tiessen et al. 1994),
10-70% in lake sediments, and 40-90% in
Histosols (Reddy, et al.!998a). Most of the
organic P in soils is derived from plant detri-
tus and synthesized in part by soil organisms
from inorganic sources (Sanyal and DeDatta
1991). Organic P can be classified into three
groups: (i) inositol phosphates, (ii) nucleic ac-
ids, and (iii) phospholipids (Anderson 1980),
with inositol phosphates comprising up to
60% of the total soil organic P (Tate 1984;
Turner 2002).
Water column P can be readily removed by
periphyton and algal uptake, followed by de-
position of dead biomass on the soil surface.
Periphyton communities are effective sinks in
low P systems. In treatment wetlands where P
loading is usually high, the periphyton effect
on P removal through direct assimilation may
be minimal. Similarly, vegetative uptake can
provide either short-term or long-term sink
for P depending on type of vegetation present
in the wetland. Phosphorus tied-up in detrital
plant/algal tissue is rapidly released into the
water column during decomposition. How-
ever, over long-term periods, significant por-
tions of organic P may remain in the soil as a
part of peat accumulation in wetlands. Under
anaerobic conditions, these forms of P are rel-
atively resistant to microbial breakdown, and
can be considered an important P sink.
9
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Inorganic P added to wetlands at concen-
trations considerably greater than those pres-
ent in the soil porewater diffuses into the
soil and is retained by oxides and hydroxides
of iron and aluminum in acid soils, and by
calcium carbonate in alkaline soils. In soils
dominated by Fe oxides, P can be readily
immobilized through sorption and precipita-
tion by ferric oxyhydroxide, and formation of
ferric phosphate in the oxidized zones at the
soil-water interface. In calcareous systems,
P can be precipitated as Ca mineral bound-
P, especially when pH of the floodwater is
altered diurnally by photosynthetic activi-
ties of algae. On the other hand, when the
amount of soluble P in the soil porewater is
higher than in the floodwater, the steep gra-
dients produced result, in the diffusive flux
of P from the soil into the overlying water
column. The P released into water column is
then either re-assimilated by planktonic or-
ganisms and precipitated as Ca-phosphate, or
released downstream. In calcareous systems,
co-precipitation of P with CaCO3 is a domi-
nant mechanism in immobilization of soluble
P. In the water column where pH fluctuates
diurnally, with high values during daytime
(photosynthetic activity of algae) and low
values during night (respiration by algae and
bacteria), P can be precipitated and resolubi-
lized (Diaz, et al.1994). The rate of adsorption
is controlled by soil pH and Eh, adsorptive
surface area (active iron and aluminum or
calcium carbonate), and temperature. These
regulators can be used as indicators to evalu-
ate P retention capacity of wetland soils.
Sulfur Cycle in Wetlands
Deposition
FIGURE 6: SCHEMATIC SHOWING BASIC SULFUR CYCLES
IN THE SOIL-WATER COLUMN OF A WETLAND.
10
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SULFUR
In freshwater wetlands, sulfur (S) cycling
was assumed to play a minimal role in regu-
lating decomposition and nutrient release.
The role of S in coastal wetlands is well es-
tablished. However, SO42 inputs from adja-
cent watersheds to freshwater wetlands can
increase the role of the S cycle; increased
inputs into saltwater wetlands can accelerate
the S cycle, especially sulfate reduction and
associated microbial processes. Similar to C
and N, organic S mineralization can be re-
lated to substrate quality, microbial biomass
and extracellular enzyme activity. Likewise,
SO42 reduction can be directly related to dis-
solved organic C and fatty acids.
In estuarine wetland soils and in freshwa-
ter environments with appreciable SO42" ,
microbial reduction of sulfate to S" or H2S
occurs. The characteristic "rotten egg" odor
of tidal and estuarine wetland soils is the
result of H2S production by sulfate reduc-
ers. Numerous species of bacteria, including
Desulfovibrio, Desulfobacter, Desulfococcus
and others, can use SO42~ as a terminal elec-
tron acceptor (Ponnamperuma 1972; Wid-
del and Hansen 1992). In wetland soils and
sediments, most SO42" reducing bacteria are
mesophiles, existing at optimum tempera-
tures of 20 to 40°C (Wieder and Lang 1988;
Widdel and Hansen 1992). Like iron (Fe) and
manganese (Mn) reduction, SO42" reduction
may be inhibited by the presence of other ter-
minal electron acceptors. Appreciable SO42
reduction does not occur in the presence of
oxidized Fe and Mn or NO3 (Ponnamperuma
1972; Yoshida 1975) because Fe/Mn reducers
and denitrifiers maintain the concentration of
electron donors at concentrations that are too
low to support populations of SO42~ reducers
(Lovley 1991).
A major difference between freshwater and
estuarine wetlands is the predominance of
sulfate reduction over methanogenesis in es-
tuarine wetlands. The concentration of SO42
in most freshwater ecosystems is only 0.1-0.2
mM as compared to 20-30 mM in seawater
(Capone and Kiene 1988). As such, increased
SO42~ availability in estuarine wetlands leads
to higher rates of sulfate reduction and addi-
tions of SO42~ to freshwater systems can change
the dominant organic matter decomposition
pathway from methanogenesis to sulfate re-
duction. (Castro and Dierberg 1987; Capone
and Kiene 1988). Estimates of sulfate reduc-
tion in salt marsh sediments are on the order
of 1000-2000 g S/m2/yr with approximately
50-90% of the total organic matter decompo-
sition occurring via this pathway (Howarth
1984; Howes, et al.1984; Howarth and Giblin
1983). In freshwater wetlands, sulfate reduc-
tion is generally limited by the availability of
SO42- (Nedwell 1984). Although sulfate re-
duction generally is lower in freshwater wet-
lands than in estuarine wetlands, significant
sulfur inputs from acid deposition may lead
to rates of sulfate reduction that are compa-
rable to those measured in estuarine soils and
sediments.
REFERENCE WETLANDS
Jo determine the nutrient impacts on wet-
lands, one must establish the background
conditions in a wetland that is not impacted
by nutrients. In addition to nutrients, impacts
on wetlands occur from hydrologic fluctua-
tions, fire, and management practices im-
11
-------
posed on these systems. All these impacts
can also influence the nutrient profiles in
wetlands. Thus, it is critical to determine the
background levels of biogeochemical indica-
tors and processes that can be used to deter-
mine the change in wetland conditions. Wet-
land sites monitored for background levels
will function as "reference" sites, which can
be used to establish wetland condition. Refer-
ence sites for sampling and monitoring can
be identified based on the historical infor-
mation available for the site. One reference
site for a given wetland type within the same
watershed may be adequate. In many wet-
lands, it may be impossible to find sites that
are not affected by anthropogenic impacts.
Under these conditions, minimally impacted
sites can be used as "reference" sites. Once
the reference sites are established, spatial and
temporal variability in selected indicators
should be monitored to determine the ranges
in values. This initial database is essential to
establish nutrient criteria.
In certain watersheds, the whole wetland
may be impacted and reference sites may not
be available. These wetlands typically have
high accumulations of organic matter. For
these sites, native soil nutrient content can be
used as background condition. This can be
accomplished by taking intact soil cores and
determining the nutrient profiles. Typically,
impacted wetlands will have high nutrient lev-
els in surface layers and decrease with depth,
and reach steady levels at lower soil depths.
Nutrient levels in lower soil depths can be
used as an indication of background levels for
that site. Depth of nutrient impact can also be
estimated by determining the age of the ma-
terial using Cs-137 dating techniques (Reddy,
et al.1993; Ritchie and McHenry 1990).
FIGURE 7: SCHEMATIC SHOWING
THE PROPOSED STEPS TO DEVELOP
AND EVALUATE BIOGEOCHEMICAL
INDICATOR CRITERIA
SAMPLING PROTOCOL
Wefore an effective method for the evalu-
IJ ation of wetland biogeochemical charac-
teristics can be established, one must identify
the portion of the wetland that: (i) responds
rapidly, accurately representing the impact
of external loading; and, (iii) provides early
warning signals of declining ecosystem health.
Changes in plant community or macrofauna
structures are often slow and the system may
be severely degraded by the time these visual
changes are observed. Concentrations of wa-
ter column nutrients are often used in other
water body types and are useful to determine
the downstream effects of impacted wetlands.
However, one prominent feature of wetland
12
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ecosystems is that water levels often fluctuate
and some wetlands have little or no period of
inundation. In addition, due to the nature of
hydrologic inputs, nutrient concentrations in
the water column of wetlands can change rap-
idly and can be highly variable. This makes
sampling of water column indicators some-
what unpredictable and requires more inten-
sive sampling to quantify indicator values.
Therefore, water column indices are useful to
evaluate input and outflow conditions from a
wetland, but do not provide the best indicator
of overall nutrient condition within a wetland.
However, water column nutrients are in di-
rect contact with the microbial communities
associated with periphyton, plant detritus,
B'l SmaJ Non-riverine
• c - cenkT-7one.
I fc - edge-zone
ci *°rei fc-composite
samples; collected within
each zone.
C) Laige Non-riverine
FIGURE 8: AN EXAMPLE OF THE SAMPLING SCHEME
USED TO PARTIALLY QUANTIFY WITHIN WETLAND
BIOGEOCHEMICAL GRADIENTS WHILE MINIMIZING WITHIN
ZONE VARIABILITY AND NUMBER OF SAMPLES FOR ANALYSIS
13
-------
and surficial soils. Thus, changes in microbial
community composition and activity, as well
as the composition of plant detritus and soils,
provide an indication of recent impact from
added nutrients.
Plant detritus and soil components func-
tion as major storages of essential nutrients
and serve as integrators of nutrient impacts.
This integration is mediated in two ways: (i)
direct assimilation of nutrients by microbes
and algae colonizing detrital plant tissues
in the water column, and (ii) assimilation of
nutrients by plant communities and deposi-
tion of enriched detritus on the soil surface.
. It should be recognized that microbes are
dependent on organic substrates provided by
macrophytes, while macrophytes are depen-
dent on microbes to transform organic forms
of nutrients in more bioavailable forms. This
mutual dependency between microbes and
macrophytes is one of the key regulators
of biogeochemical processes in wetlands.
Therefore, biogeochemical processes and as-
sociated microbial communities that respond
rapidly to nutrient loading and related physi-
co-chemical properties of soil, detritus, and
water column, yet integrate the high variabil-
ity of nutrient loading often associated with
pulsed runoff events, can function as useful
indicators to determine nutrient impacts.
Nutrient inputs to wetland can occur at var-
ious points from point and non-point sources.
The effects of nutrient loading are usually
patterned with the impacted zone adjacent to
inflow points and the unimpacted zones in the
interior of a wetland. Thus, monitoring sta-
tions should be located in both impacted and
unimpacted zones to quantify nutrient loads
accurately. Biogeochemical indicator selec-
tion and evaluation requires systematic steps
before incorporation into a routine monitoring
program (Figure 7). Module 4: Study Design
for Monitoring Wetlands, provides a thorough
discussion on study design and sampling
protocols depending on wetland class and
hydrology.
One example of a sampling design used to
survey wetlands throughout the southeast-
ern United States implemented a two zone
composite sampling scheme. This technique
allows for limited evaluation of wetland bio-
geochemical gradients due to source of nu-
trient inputs and hydrologic effects. It mini-
mizes variability as well as the total number
of samples to be analyzed. This sampling
scheme can be applied to riparian as well as
non-riparian wetlands. In both wetland types,
the process begins with a visual survey of the
wetland upon arrival and then divides it into
two general zones, referred to as the center-
zone "c" and the edge-zone "e" (Figure 8). In
riverine systems, the center-zone is adjacent
to the stream, but landward of any natural
levees that may have formed. The edge-zone
of riverine wetlands is located parallel to the
upland, approximately one-third the distance
between upland and stream. In non-riverine
wetlands, a similar zone criteria can be ap-
plied where the center third of the wetland is
designated as the center-zone and the driest
third of the wetland is designated the edge-
zone.
Subsampling within the center and edge
zones varied slightly depending on wetland
size but generally, consisted of a triplicate
composite sample of soil, litter and water at
each sampling station. In riverine wetlands
and large (> 10 ha) non-riverine wetlands,
samples were collected along one side of
the wetland by collecting three subsamples
14
-------
25 paces apart along a transect parallel to
the wetland topographic contour. In small
(<10 ha) non-riverine wetlands, samples were
collected within edge and center zones at
three equidistant points around the wetland.
Using this method, samples at each wetland
can be collected in approximately one hour.
One sample for each zone and strata is ana-
lyzed and used to characterize the wetland.
Although information about within-zone
variability is not retained due to compositing,
greater confidence in the true central tenden-
cy of values is provided without having to run
three times as many samples in the labora-
tory. When conducting a broad baseline sur-
vey for reference or monitoring purposes this
method may provide a suitable compromise
between information gained and available re-
sources.
WATER SAMPLING
TT/ater quality monitoring protocols to as-
F r sess trophic conditions for natural wet-
lands have not been established. However,
wetlands used for treating wastewaters are
heavily monitored, and the protocols required
to monitor these systems are well established.
The objective of monitoring wastewater
treatment wetlands is to evaluate the nutri-
ent removal efficiency. These wetlands are
typically monitored at the inflow and outflow
points for various water quality parameters.
Similarly, water-sampling stations in natu-
ral wetlands should be established to capture
the spatial heterogeneity caused by nutrient
inputs, vegetation and periphyton communi-
ties. At the minimum, water samples should
be obtained at inflow and outflow points and
selected stations within wetlands. In addi-
tion, the sampling frequency should capture
temporal variability within parameters being
monitored. Rainfall and hydraulic loading
affect water depth and concentration of vari-
ous water column constituents. It also influ-
ences biotic communities. Thus, it is critical
to measure water depth and stage in conjunc-
tion with selected water quality parameters
at all water-sampling stations. Water samples
should be collected using the protocols ap-
proved by APHA (1992) and USEPA (1993).
SOIL SAMPLING
TJ/etlands exhibit a high degree of spatial
f r heterogeneity in chemical composition
of detrital and soil layers. Areas impacted by
nutrients may exhibit a high degree of vari-
ability as compared to unimpacted sites with-
in the same wetland. Thus, sampling proto-
cols should capture this spatial variability.
Developing and monitoring successful nutri-
ent criteria management programs require
sampling protocols that to capture spatial and
temporal patterns.
Soil samples are usually obtained using ei-
ther a grab sampling approach or collection
of intact soil cores. Grab sampling is not suit-
able for characterizing soils, because it does
not provide any indication of soil bulk density.
Grab samples obtained from a constant depth
(example: 0-5 cm or 0-10 cm) can be used to
determine the nutrient enrichment in surface
soils. However, comparison among different
soils types is difficult, if intact soil cores are
not obtained to determine nutrient concentra-
tions and bulk density. For example, nutrient
concentrations (expressed per unit dry weight
of soil) in wetlands with light soils (such as
organic soils) will be high in nutrients as
15
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TABLE 1: POTENTIAL WATER AND SOIL QUALITY INDICATORS
FOR ASSESSING NUTRIENT IMPACTS IN WETLANDS.
CDENOTES MINIMUM DATA REQUIRED FOR EACH SITE).
Color*
Temperature*
Water depth*
Salinity/Conductivity*
Turbidity*
To tal suspended solids*
Dissolved oxygen
pHand alkalinity*
Nitrogen (TKN)*
Phosphorus (TP)*
Soil bulk density*
Soil pH and Eh*
Cation exchange capacity
To tal N, and P*
Organic matter content*
To tal carbon, and labile carbon
Particle size distribution
C:N:P ratios*
Extractable nutrients*
Sediment oxygen demand
Acid volatile sulfides
Oxalate extractable metals
Nitrogen (NH4-N, NO 3 + NO2-N)
Phosph orus (TDP , and DRP)
To tal and dissolved metals (site specific situation)
To tal and dissolved organic C (TO C and DO C)
To tal and dissolved organic N (TO N and DO N)
Elemental ratios (Si:C:N:P)
Enzyme assays
Heterotrophic respiration
UV absorbence
Biological N2 fixation
Periphyton community composition
Enzyme assays
Organic matter/soil accretion rates
Stable isotope ratios
Soil-water nutrient exchange rates
Cotton-strip assay
Detrital decompo sition —litter bag
Microbial activity- Respiration
Microbial activity- Methanogenesis
Microbial biomass C, N, and P
- Organic N and P m ineralization
Nitrification
D enitrification
Sulfate reduction
Substrate induced respiration
Arginine mineralization
Pho spho rus sorp tion /desorp tion;
Partition coefficients for P
- P Adsorption maximum
Degree of P saturation
Soil mineralogical compo sition
Microbial diversity
compared to wetlands with heavy soils (such
as mineral soils), even though both wetlands
may have similar impacts. This problem can
be corrected by expressing soil nutrient con-
centrations per unit volume of soil, which re-
quires the measurement of soil bulk density.
16
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Since intact soil cores are typically used
to characterize wetland soils, caution must
be used to ensure that the soil is minimally
disturbed when obtaining the core samples.
Several approaches have been used to obtain
undisturbed soil cores from wetlands. These
include the use of PVC, acrylic, and alumi-
num tubing as coring devices. Core diameter
is critical to avoid compaction. Coring tubes
with diameter of <10 cm cause considerably
more soil compaction than 12-15 cm tube
diameters. Standard coring probes used in
upland soils are not suitable for wetlands,
because of saturated soil conditions and low
bulk densities. Typically, organic-rich wet-
land soils have bulk densities in the range of
0.1-0.3 g cm"3 (dry weight). Nutrient concen-
trations expressed both on mass and volume
basis should be used in evaluating the degree
of impacts. However, bulk density measure-
ments are often less precise and potentially
may add an additional source of variation.
Intact soil cores with little or no detectable
soil compaction can be obtained using an alu-
minum cylinder (15 cm diameter), with sharp-
ened lower edge, that can be twisted through
the fibrous marsh soils to a depth of 60 cm.
The top of the cylinder is sealed with a PVC
cap or a stopper to provide suction, and the
bottom of the cylinder is sealed with a rubber
stopper. The intact cores are then removed
from the soil and sectioned into desired depth
increments. The surface detritus is removed
from the soil and saved for chemical analysis.
Typically, soil cores are sectioned into 0-10,
10-30, and 30-60 cm for routine character-
ization. Selection of depth increments should
be based on site-specific conditions and soil
profile characterization. The approach de-
scribed above has been used in several stud-
ies by researchers at the University of Florida
and Louisiana State University. For routine
monitoring of soil properties, typical root
zone depth (0-10 and 10-30 cm) may ad-
equately characterize the system.
BlOGEOCHEMICAL
INDICATORS
everal biogeochemical indicators can be
used as response variables to evaluate nu-
trient impacts in the soil and water of a wet-
land. Level I soil and water quality indicators
are relatively easy to measure, and many are
now routinely used either in monitoring aquat-
ic systems, such as streams, rivers, lakes, and
estuaries or terrestrial ecosystems. Level II
indicators are relatively complex measure-
ments, although many of these indicators are
now measured in wetland systems. Level II
indicators provide a better understanding of
the influence of nutrient enrichment on soil
processes and its ultimate effect on ecosys-
tems function. However, both processes and
indicators are regulated by various controls
including nutrient loads, hydrology, fire,, and
spatial/ temporal variability. Many of the
Level I water and soil quality indicators can
be used as independent response variables
that may influence the biogeochemical pro-
cesses functioning in the soil and water of a
wetland. As the dependent variable, biogeo-
chemical processes can be used as key re-
sponse variables affected by nutrient loading.
A list of Level I and II indicators are shown
in Table 1.
17
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WATER QUALITY INDICATORS
TEMPERATURE, DISSOLVED
OXYGEN AND PH
utrophication, as defined, refers to an
overabundance of nutrients (nitrogen
and/or phosphorus) in a wetland ecosystem
that produces adverse water quality impacts.
In addition to nutrients, anthropogenic loads
of organic matter, suspended solids, trace
metals, and pesticides can also impair water
quality and associated biotic communities.
A number of physical, chemical, and biologi-
cal water quality parameters are now used as
indicators of water-body impairment. Water
quality variables should be evaluated criti-
cally to obtain the most cost-effective infor-
mation required to assess wetland impairment.
Specifically, water quality monitoring should
determine the range of values that would sig-
nificantly impair the ecological integrity of
wetland. The following list of variables may
help address the questions related to ecologi-
cal impacts on wetlands. The significance
of these variables is clearly established for
streams, lakes, and estuaries. Once the data
are obtained, the relationships between water
quality variable and response variables should
be established. For example, the relationship
between nutrient concentrations and response
variables such as algae, vegetation, and ben-
thic invertebrates are useful when evaluating
impacts on wetlands. However, sampling of
the water column is restricted to periods when
the wetland is flooded. Although, the useful-
ness of water quality data is restricted to cer-
tain time of the year, it may provide useful
information regarding short-term temporal ef-
fects on wetland biota resulting from nutrient
loading. Detailed methodologies for determi-
nation of water quality parameters can be ob-
tained from following sources: APHA (1992);
Clesceri, et al., (1998), and Wetzel and Likens
(1990).
Water temperature controls many of the mi-
crobially mediated biogeochemical reactions
in the water column. Variations in tempera-
tures are reflected in the ranges of values for
various water quality parameters and in the
productivity of periphyton and vegetation.
The effects of temperature on biogeochemical
processes are well documented in the litera-
ture, with many of the reactions responding to
temporal patterns in water temperature. In ad-
dition to distinct seasonal patterns, these tem-
poral patterns in temperature can be observed
within a diurnal cycle. However, amplitude of
the daily water temperature swing depends on
local climate and wetland type. Variations in
temperature can be much higher in marshes
with emergent vegetation, as compared to for-
ested wetlands. Water temperature can be eas-
ily monitored with very inexpensive instru-
ments.
Concentration of dissolved oxygen (DO) in
the water column readily responds to anthro-
pogenic impacts. Highly degraded wetlands
may have wide shifts in DO concentrations.
For example, wetlands receiving waters con-
taining carbonaceous and nitrogenous oxygen
demand can exhibit oxygen depletion in the
water column. Oxygen production by algae
can increase daytime DO concentrations and
may result in low DO concentrations during
the night. Among the Level I indicators, DO
is and easily measurable, and commonly used
by agencies involved in managing aquatic eco-
systems. Oxygen is consumed during biologi-
cal and chemical processes functioning in the
water column. Plant, animal, and microorgan-
isms consume oxygen during respiration. Sim-
ilarly, nitrification (oxidation of ammonium to
18
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nitrate and nitrite N) and oxidation of reduced
substances such as sulfides, methane, and re-
duced iron and manganese consume oxygen.
Excessive oxygen production in the water
column leading to super-saturation levels dur-
ing daylight periods is often an indicator of
nutrient enrichment in open canopy wetlands.
At night, hypoxia can result from increases
in detrital production, heterotrophic respira-
tion, and release of reduced compounds at
the same location. Oxygen is measured easily
using commercially available oxygen probes.
Single point measurements of oxygen may not
be meaningful and may have very little value
in evaluating impacts on wetlands. However,
monitoring of diurnal fluctuations in oxygen
levels continuously is more meaningful and
can help explain the fate of nutrients in the
water column.
The pH of the water column affects many
biogeochemical processes. Water column pH
can be highly variable depending on wetland
type. Typically, the pH of water in many om-
brotrophic wetlands (wetlands that predomi-
nantly receive hydrologic inputs from rain-
fall) is acidic, while pH of eutrophic wetlands
is much closer to neutral. Photosynthetic ac-
tivity of algae and submersed aquatic vegeta-
tion influences water column pH in poorly
buffered waters. Photosynthesis results in de-
pletion of CO2 in the water column, shifting
the carbon dioxide - bicarbonate - carbonate
equilibria. However, during the night, high
rates of respiration increase the production of
protons, thus resulting in decrease in pH of
the water column. Similar to oxygen moni-
toring, single point measurements of pH may
not be useful in evaluating impacts on wet-
lands. Thus, continuous recording probes are
useful in monitoring pH in the water column.
SUSPENDED SOLIDS,
TURBIDITY, AND COLOR
Wetlands are effective when removing sus-
pended solids and turbidity from inflow wa-
ters, due to lower water velocities and veg-
etation, which promote filtration and rapid
physical settling of solids. Settling of sus-
pended solids occurs in areas closest to the
inflow point in wetlands. Many pollutants
including nutrients, metals, and toxic organ-
ics are associated with suspended solids. Sus-
pended solids often retain contaminants and
can therefore maintain low concentrations of
dissolved contaminants in the water column.
Sampling the suspended solids in the water
column within the interior marsh is often
challenging, because of disturbance associ-
ated with acquiring the sample in a shallow
water column and solids production within
the wetland. Internal generation of solids oc-
curs through fragmentation of detritus and
litter, algal cells, and bioturbation by benthic
invertebrates further contributing to the con-
centration of total suspended solids (TSS) in
a wetland. Total suspended solids are easily
measured. Water samples are acquired, fil-
tered and dried, then weighed to determine
the concentration of suspended solids (APHA
1992). In situ probes for measuring TSS are
also available.
The primary causes of turbidity in water are
suspended solids and color. Turbidity is mea-
sured using a turbidimeter that consists of a
nephelometer, light source, and photodetector
(for details see APHA 1992, EPA, 1993). The
turbidity standard unit is expressed as NTU
(nephelometric turbidity unit). The following
relationship between turbidity and total sus-
pended solids (has been reported by Kadlec
and Knight (1995):
19
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NTU = 0.83TSS; R2 = 0.77
Color is a qualitative water quality param-
eter. Color in surface waters of wetlands may
be derived from decomposition of organic
matter, plankton and phytomass, and natu-
rally occurring metallic cations. Presence of
turbidity in the water interferes with the mea-
surement of color, thus some level of pretreat-
ment is necessary to remove suspended sol-
ids. The color of surface waters is measured
by both visual and spectroscopic methods
(APHA 1992).
HARDNESS
Total hardness refers to the sum concen-
tration of divalent cations such calcium and
magnesium, both expressed as CaCO3, in
milligrams per liter. Surface water with low
hardness is referred to as soft water, contain-
ing low concentration of calcium and magne-
sium. For example, rainwater is considered as
soft water with calcium concentrations in the
range of 0.1 to 10 mg/L and magnesium con-
centrations of 0.1 mg/L and a hardness value
of <30 mg/L as CaCO3. Many ombrotrophic
wetlands and bogs typically have low base
cations and can be grouped as soft water wet-
lands. These wetlands are typically acidic in
nature. Methods to determine the hardness of
water are described in APHA (1992). Hard-
ness of water is calculated as follows:
Hardness, mg equivalent CaCO3/L =2.497
[Ca, mg/L] + 4.118 [Mg, mg/L].
CONDUCTIVITY
Conductivity (also referred to as specific
conductance or electrical conductivity) refers
to the activity of total dissolved solids and
their ability to conduct electrical current. Sur-
face waters with a high concentration of inor-
ganic compounds provide high conductivity.
Conductance is defined as the reciprocal of
resistance and is expressed as |imhos/cm. In
the international system of units (SI) conduc-
tivity is reported as milliSiemens per cm (mS/
cm); 1 mS/cm = 10 |imhos/cm (APHA 1992).
The conductivity of most natural freshwaters
is in the range of 1 to 30 mS/cm, while con-
ductivity values may reach levels >6,000 mS/
cm in depressional salt lakes. The conductiv-
ity of surface water in deepwater swamps is
in the range of 6-55 mS/cm (Mitsch and Gos-
selink 1993). Conductivity is usually mea-
sured using in situ probes (APHA 2000); ad-
ditional information about conductivity and
its measurement are found in APHA (2000).
SALINITY
Salinity is a measure of total dissolved ion
concentration of water and is used as a ref-
erence to estuarine and marine environment
including salt marshes, mangrove wetlands,
and estuaries. Salinity is reported as parts
per thousand (ppt) or one gram of salt in one
kg of water (APHA 1992). Salinity is usually
measured using an algorithm based on tem-
perature and electrical conductivity, density,
or light refraction (APHA 2000).
NITROGEN [TKN, NH4-N,
AND NO3 + NO2 -N]
Nitrogen (N) is present in organic and in-
organic forms in surface waters. Methodolo-
gies to monitor surface waters are well devel-
oped for other ecosystems and can be readily
adopted for wetlands. The most commonly
monitored N species are total Kjeldahl ni-
trogen (TKN), ammonium N (NH4-N), and
20
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nitrate plus nitrite N (NO3 + NO2-N) (APHA
1992; EPA 1993). The TKN analysis includes
both organic and ammonium N, but does not
include nitrate plus nitrite N. Organic N is de-
termined as the difference between TKN and
NH4-N. Nitrogen concentration in the water
column is typically expressed as mg N/L.
Wetlands vary considerably in their capac-
ity to process and assimilate N. For example,
wetlands are very effective in removing nitrate
N through denitrification process, but may
not be as effective in removing ammonium N
and organic N. Internally wetlands produce
soluble organic N through decomposition of
organic matter. Nitrogen removal efficiency
is often determined by the hydraulic retention
time, wetland type (including variations in
vegetation and soil type), and climate. Thus,
the establishment of critical range values for
N needs to include the variability among wet-
land community types within a given region.
PHOSPHORUS
[TP, TOP, AND DRP]
Phosphorus (P) entering a wetland is typi-
cally present in both organic and inorganic
forms. The relative proportion of each form
depends upon soil, vegetation and land use
characteristics of the drainage basin. To trace
the transport and transformations of P, it is
convenient to classify P entering into these
systems as: (i) dissolved inorganic P (DIP);
(ii) dissolved organic P (DOP); (iii) particu-
late inorganic P (PIP); and, (iv) particulate
organic P (POP). The particulate and solu-
ble organic fractions may be further sepa-
rated into labile and refractory components.
Dissolved inorganic P is considered bioavail-
able, whereas organic and particulate P forms
generally must undergo transformations to
inorganic forms before being considered bio-
available. Phosphorus retention in wetlands is
regulated by many factors including vegeta-
tion, periphyton and plankton, plant litter and
detrital accumulation, soil physico-chemical
properties water flow velocity, water depth,
hydraulic retention time, length to width ra-
tio of the wetland, P loading, and hydrologic
fluctuations. When evaluating wetlands for
P assimilation and establishment of critical
ranges, it is necessary to consider: (i) short-
term storage mediated by assimilation into
vegetation, periphyton and incorporation into
detrital tissue; and, (ii) long-term storage me-
diated by soil assimilation, and accretion of
organic matter.
Phosphorus concentration in the water col-
umn is typically expressed as |ig/L or mg/L.
Methodologies to monitor P in surface waters
are well developed for other ecosystems and
can be readily adopted for wetlands. The most
commonly monitored for P species are total
P (TP), dissolved reactive P (DRP), and total
dissolved P (TOP) (APHA 1992; EPA 1993).
ALKALINITY
Alkalinity of water refers to its acid neutral-
izing capacity and is primarily a function of
carbonate, bicarbonate, and hydroxide levels.
It is expressed as mg CaCO3/L. A number of
wet chemistry methods are available to deter-
mine alkalinity of surface waters in wetlands
(APHA 1992).
21
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TOTAL, DISSOLVED
AND PARTICULATE
ORGANIC CARBON
Total organic carbon (TOC) in surface wa-
ters includes: dissolved organic carbon (DOC) —
the fraction of TOC that passes through a
0.45 um pore diameter filter, and particulate
organic carbon (POC) — the fraction of TOC
that is retained on the 0.45 um diameter fil-
ter. Since some of the organic matter in sur-
face water can be oxidized or utilized by mi-
crobes, simple methods such as biochemical
oxygen demand (BOD) and chemical oxygen
demand (COD) are often used as indicators of
organic carbon (APHA 1992). However, total
organic carbon may contain biodegradable
and recalcitrant organic matter. Thus, direct
measurement of TOC may be more useful in
determining the water column organic car-
bon content. A number of methods such as
direction combustion at high temperature or
combustion using infrared method, or per-
sulfate-ultraviolet oxidation method are used
to determine TOC in surface waters (APHA
1992). The TOC concentrations can also be
related to the turbidity or color of the surface
waters. Dissolved organic C, TOC, and POC
concentrations are expressed as mg/L.
SULFATE AND SULFIDE
may not be critical for water quality. Howev-
er, recent linkages between sulfate reduction
and mercury methylation in wetlands sug-
gest that monitoring sulfates for site-specific
conditions may be essential. Accumulation
of sulfides in the soil may have positive ef-
fects on availability of metals, as sulfides can
precipitate metals into insoluble forms. This
process is very important in coastal wetlands
where sulfate concentrations are several fold
higher than freshwater wetlands.
Sulfate concentrations are expressed in
mg/L. Sulfate in surface waters can be ana-
lyzed using ion chromatographic method or
gravimetric method (APHA 1992). Sulfides
can be measured using a sulfide electrode or
by wet chemistry methods (APHA 1992).
METALS
Monitoring wetlands for metals can be im-
portant under site-specific conditions. Met-
als can be determined using several meth-
ods including atomic absorption, inductively
coupled plasma (ICP), or using wet chemistry
methods (APHA 1992). These methods are
widely used by both governmental and com-
mercial laboratories. Metal concentrations in
the water column are expressed as mg/L or
In the absence of oxygen, nitrate, and oxi-
dized forms of iron and manganese, obligate
anaerobic bacteria can use sulfate as their
electron acceptor during respiration, while
using organic matter as their energy source.
Under most wetland conditions, sulfates in
the water column are stable. However, sul-
fates can diffuse from water column to un-
derlying soil, where they can be reduced to
sulfides. In many freshwater wetlands, sulfate
concentrations are low, so monitoring them
Dissolved metals are those constituents
(metals) of a sample that pass through 0.45
um filter and are preserved under acid condi-
tions (pH <2). Acid extractable metals repre-
sent the concentration of metals in the sample
after treatment of unfiltered sample with hot
dilute mineral acid. Total metals are those de-
termined in unfiltered sample after digestion
with acids. Several digestion methods are
routinely used (APHA 1992).
22
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SOIL QUALITY INDICATORS
Ooil quality may be viewed as the capacity
kJ of the soil to function within ecosystem
boundaries and to sustain biological produc-
tivity, maintain environmental quality, and
promote plant and animal health. Soil in-
cludes both native soil and detrital/litter com-
ponents which potentially convert into soil
organic matter. A number of soil and detrital
physical, chemical, and microbial parameters
can be used as potential indicators of qual-
ity. We will use soil indicators that directly
or indirectly reflect quality, and can be linked
with environmental or anthropogenic fac-
tors. For example, dissolved constituents and
those bound on soils contribute significantly
to surface water quality. Chemical and mi-
crobial indicators can also provide important
links to soil quality. This class of ecological
indicators is associated with biogeochemical
cycling, the turnover and storage of nutrients
and other elements in biotic and abiotic com-
partments of the ecosystem. Therefore, it is
related to biological productivity in the soil
as well as in the floral and faunal components
of the ecosystem. Recent studies in terrestrial
and wetland ecosystems have demonstrated
the potential of using microbial communities
and process measurements as sensitive indi-
cators of environmental perturbations in soils
and sediments (Torstensson, et all998), in-
cluding heavy metals (Frostegard, et al.1993),
physical disturbance (Findlay, et al.1990), and
nutrient impacts (DeBusk and Reddy 1998;
Reddy, et al. 1999).
Decomposition of organic matter is the pri-
mary ecological role of heterotrophic micro-
flora in soils. It provides for mineralization
of potentially growth-limiting nutrients and
formation of recalcitrant organic compounds
(e.g. humus) that contribute to the chemical
stability of the system (Swift 1982; Jorgens-
en, et al. 1999; Middelboe, et al. 1998). Soil
microbes may also exert a significant influ-
ence on ecosystem energy flow in the form
of feedback, since the mineralization of or-
ganically bound nutrients is a regulator of
nutrient availability for both primary produc-
tion and decomposition. Therefore, most of
the net ecosystem production passes through
the microbial compartment at least once and
typically several times, although microbial
biomass comprises only a small fraction of
the sediment organic matter (Jorgensen, et
al. 1999; Ruttenberg and Goni 1997; Seitz-
inger and Sanders 1997). Many of the process
level parameters are tedious to measure and
require specialized training and equipment;
thus, they may not be suitable for routine
monitoring in wetlands.
Monitoring soil properties may provide
long-term integrated effects of nutrient im-
pacts on wetlands, but may not be suitable to
determine short-term temporal changes in the
system. Chemical composition of soils and
detrital components of wetlands can be easily
monitored using approaches similar to that
used in soil testing for plant nutrition in agri-
cultural ecosystems. However, some modifi-
cations are necessary to adopt these methods
to wetland environment. Detailed methodolo-
gies for determination of soil quality param-
eters can be obtained from following sources:
Dane and Topp, (2002), Klute (1986), Sparks,
et al. (1996), Weaver, et al. (1994).
SOIL BULK DENSITY
Soil bulk density refers to the ratio of the
mass of dry solids to the bulk volume of soil
(Blake and Hartge 1996). The bulk volume
23
-------
includes the volume of solids, water, and pore
space. The mass of solids is determined after
drying the soil at 105°C and the bulk density
is calculated as follows:
Bulk density (dry) (g/cm3) = (mass dry
weight, grams)/volume (cm3)
Bulk density of wetland organic soils rang-
es from 0.1 to 0.4 g/cm3, while for mineral
wetland soils values can be in the range of 1
to 1.5 g/cm3. This is a useful parameter, when
concentration of nutrients is expressed on a
volume basis rather than mass basis. For ex-
ample, concentration of nutrients in organic
wetland soils can be high when expressed as
mg/kg of dry soil (|ig/g of dry soil), as com-
pared to mineral wetland soils. However,
the difference in concentration may not be
as high when expressed on a volume basis.
Nutrient availability to plants is regulated by
the total amount of nutrients available in soil
volume. Bulk density is determined on intact
soil cores obtained at known depth and vol-
ume. Volume is determined from core diame-
ter and depth, while the dry weight of the soil
per core volume is determined as described
above.
SOIL PARTICLE SIZE
DISTRIBUTION
Particle size analysis (PSA) is a measure-
ment of size distribution of individual par-
ticles in a soil sample (Gee and Bauder 1986).
Particle size analysis is only applicable to
mineral wetland soils. Soil particles cover a
wide range including clay (<0.002 mm), silt
(0.002 to 0.02 mm), fine sand (0.02 to 0.25
mm), coarse sand (0.25 to 2 mm), and gravel
(2 to 80 mm). Particle size analysis is often
used to determine soil texture, which is based
on various combinations of sand, silt , and
clay separates that make up the particle size
distribution of a soil sample (Gee and Bauder
1986). Particle size distribution also corre-
lates to the phosphorus and metal retention
capacity of soils. Typically the larger the clay
and silt fraction of a sample the greater ca-
pacity the soil has for P and metals retention;
however, this is also highly dependant upon
the mineralogy of these smaller size frac-
tions. Several simple methods are described
by Gee and Bauder (1986) for PSA of mineral
upland soils, which can be readily applied to
wetland soils.
REDOX POTENTIAL
Oxidation and reduction reactions regulate
many of the biogeochemical reactions in wet-
lands. Electrons are essential to many bio-
geochemical reactions. Oxidation, the loss of
electrons, couples with reduction in the gain
of electrons. The Eh of soil is determined by
the concentration of oxidants and reductants.
Oxidants include oxygen, nitrate, nitrite,
manganic manganese, ferric iron, sulfate and
CO2, while reductants include various organic
substrates and reduced inorganic compounds.
From the ecosystem point of view, photo-
synthesis and respiration are two examples
of reduction and oxidation reactions, which
regulate energy flow and many biogeochemi-
cal reactions. For example, during photosyn-
thesis, CO2 is reduced to carbohydrate, and
during respiration the reduced organic com-
pounds are oxidized to CO2. Photosynthesis
provides an organic matter source to the soil
through plant productivity and detrital accu-
mulation. Compared to upland soils, one of
the most striking characteristics of a wetland
soil is its low Eh, which is a measure of elec-
24
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tron activity or potential in the soil. Redox
potential of soil is measured using a platinum
electrode with a standard calomel reference.
Redox potential (Eh) measurements have
been widely used to characterize the pro-
pensity of wetland soils to oxidize or reduce
substances. The value of these measurements
depends on their interpretation with due rec-
ognition to its theoretical and practical limi-
tations. Redox potential is the best available
simple indicator of the oxidation-reduction
status of soil, for the following reasons: (i)
the range of Eh values in wetland soil is much
wider, approximately 1000 mV, than the 300
mV range of Eh values in drained soils; (ii)
the higher concentrations of reductants that
contributes to the mixed potential in wetland
soils result in better poising and better re-
producibility; (iii) O2 is easily reduced and,
therefore, usually present at negligible levels
in wetland soils. Methods used to determine
O2 status of drained soils cannot be used in
wetland soils; and, direct measurement of
reducing compounds (i.e., the presence of
ferrous iron, manganese, or sulfides or meth-
ane) is cumbersome and difficult to perform
on a routine basis. Thus, the redox potential
remains a generally applicable, reasonably
convenient way to identify the presence and
intensity of reduction in wetland soils.
Redox potential is significantly affected
by the oxygen content of the soil. In systems
where aerobic organisms function, the Eh
range is very narrow, ranging from approxi-
mately +700 mV down to about +300 mV. Be-
low values of 300 mV, facultative anaerobes
function down to about 0 mV. Below this
range, obligate anaerobes function. In wet-
land (waterlogged or flooded) soils, Eh can be
anywhere along the entire scale. Where oxy-
gen is present in wetland soil the Eh can be as
high as in a drained soil, but where oxygen is
not present Eh can be low (-250 to -300 mV).
Redox potentials in soils are measured in:
(i) soil porewater; (ii) soil slurry; (iii) intact
soil cores; or, (iv) in situ under field condition.
Redox potential is measured using a platinum
electrode (custom made or purchased com-
mercially) and reference electrode (usually
calomel electrode), both inserted in the soil
or soil slurry. Both electrodes are connected
to a pH/millivolt (mV) meter. Readings sta-
bilize within minutes; however, it is recom-
mended that electrodes be left for at least 24
hours before measurements are taken in the
field. A longer equilibration period is neces-
sary because of the heterogeneous nature of
soil systems. The reference electrode does
not have to be at the same depth as platinum
electrode, as long as there is enough soil
moisture to insure a good electrical contact.
However, the distance between the platinum
electrode and the reference electrode can af-
fect the electrical resistance of the measuring
circuit. This may not be a serious problem in
saturated soils. For details on methodologies,
see Patrick et al., 1996.
There are several commercially available
electrodes designed to measure Eh, pH and
specific ion activities. This section will pres-
ent simple methods to construct redox elec-
trodes for use in the laboratory and under field
conditions. Many commercially available
electrodes are bulky and are not suitable for
use under field conditions. Methods associat-
ed with the construction of redox electrodes
were developed for the past three decades in
our laboratories.
25
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Field electrodes are typically made with
platinum wire with a diameter of 1.024 mm
cut into 1 to 1.5 cm segments. Smaller gauge
platinum is not recommended because of
poor reproducibility of potentials. These seg-
ments are soaked in a concentrated acid (1:1
HNO3 and HC1 mixture) for several hours
(< -4 hours) to remove surface contamina-
tion of metals and other impurities, followed
by thorough washing in deionized distilled
water. The platinum segments are soldered
or fused to a copper lead (1.628-2.053 mm
diameter) of desired length. Waterproof ep-
oxy glue is used to cover the joint between
copper and platinum. The copper wire is in-
serted into a heat shrinking tube to proved
insulation. The platinum tip end of the copper
wire is sealed with epoxy, while at the other
end, about 3 cm of copper wire is exposed.
This end is connected to an insulated cable
connected to a pH/volt meter. This type of
electrode can be left installed in the field for
a period of about one year for seasonal mea-
surements. However, it is recommended that
the electrodes are periodically checked and
reinstalled in the field (at least once every
three months).
Redox potentials or Eh values are typically
expressed in reference to a standard hydrogen
electrode (SHE). Using a SHE is not practi-
cal; therefore, alternate reference electrodes
are used. Reference electrodes (calomel elec-
trodes) are available commercially through
scientific catalogs or can be prepared in the
laboratory as follows. Calomel electrodes are
checked by substituting a standard calomel
electrode for the platinum electrode and veri-
fying a zero potential difference between the
two half-cells. Commercially available refer-
ence electrodes are convenient for use under
field conditions. To convert electrode poten-
tial meter readings to Eh values using calo-
mel reference electrodes, the meter readings
are adjusted by adding +245 mV to the read-
ings. For example, if the electrode potential
meter using calomel electrode reads -100 mV,
then the actual corrected Eh values will be:
(-100 mV + 245 mV) = 145 mV. If silver/sil-
ver-chloride reference electrodes were used,
then +199 mV is the correction factor. Note
that these correction factors are temperature
dependent. However, changes in potential due
to temperature are usually small compared to
the variability in electrode potential measure-
ments made in the field. Thus, correcting for
temperature effects on Eh values is not critical.
SOILPH
The pH of wetland soils and water varies
over a wide range. Organic wetland soils are
often acidic, and mineral wetland soils are
frequently neutral or alkaline. Flooding a soil
results in the consumption of electrons and
protons. In general, flooding acidic soils in-
creases the pH, while flooding alkaline soils
decreases pH (Mitsch and Gosselink 1993).
The increase in pH of flooded soils is largely
due to the reduction of iron and manganese
oxides. However, the initial increase in pH
can also occur due to rapid decomposition of
soil organic matter and accumulation of CO2.
The decrease in pH that generally occurs
when alkaline soils are flooded results from
the buildup of CO2 and carbonic acid. Ad-
ditionally, the pH of alkaline soils is highly
sensitive to changes in the partial pressure of
CO2. Carbonates of iron and manganese can
also buffer the pH of soil to neutrality.
Soil pH is measured using commercially
available combination electrodes on soil slur-
ries. If air dry or moist soil is used, a 1:1 soil
26
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to water ratio should be used. For details on
methodology, the reader is referred to Thomas
(1996).
LOSS ON IGNITION OR SOIL
ORGANIC MATTER
Wetlands are characterized by the accumu-
lation of organic matter resulting from high
productivity and slow rate of decomposition.
In addition to providing nutrient storage and
supply, organic matter also increases the cat-
ion exchange capacity of soils, increases the
adsorption or deactivation of organic chemi-
cals and trace metals, and improves overall
soil structure. Loss on ignition (LOT) is a
measurement of the organic matter content of
soil. A number of methods are now routinely
used to estimate organic matter content of the
soil and expressed as total organic carbon or
LOT. Significant relationships have been re-
ported between total carbon and organic mat-
ter content or LOT. The ratio of organic mat-
ter content to total carbon ranges from 1.8 to
2.2, while the ratio of LOT to total carbon was
found to be 2.6. The ratio of organic matter to
LOT was reported to be in the range of 0.57 to
1.13 for a wide range of mineral upland soils
(Nelson and Sommers 1996).
TOTAL NITROGEN AND
PHOSPHORUS
Total Kjeldahl nitrogen (TKN) in soils
and detritus/litter material is determined by
converting organic forms of N to NH4-N by
digestion with concentrated H2SO4 at tem-
peratures of 300-350°C (Bremner 1996). The
NH/rN in digested samples is analyzed using
colorimetric methods (APHA 1992). Methods
of total N determination are well defined and
are now routinely used by many laboratories.
Total P in soils and detritus/litter material
is determined by oxidation of organic constit-
uents and acid dissolution of minerals at tem-
peratures of <300°C (Kuo 1996). Digested
solutions are analyzed for P using colorimet-
ric methods (APHA 1992). Methods of total
P determination are well defined and are now
routinely used by many laboratories.
ORGANIC MATTER AND
SEDIMENT ACCRETION
Organic matter accretion refers to accumu-
lation of detrital material derived from vege-
tation and algae, while sediment accretion re-
fers to the accumulation of both organic and
inorganic material. Wetlands are effective in
trapping sediments. Sediment accretion can
be measured by placing a suitable marker,
such as feldspar, and measuring the accumu-
lation of both organic and inorganic material
over time.
Accretion rates are also determined by gam-
ma analysis of the 661 keV photopeak for 137Cs
in each increment using a low-energy germa-
nium detector (Canberra Industries, Meriden,
CT). Cesium-137 (half-life 30 yr), produced
from above-ground thermonuclear weapons
testing, can be used as a marker to estimate
recent (30-40 yr) rates of sediment deposi-
tion and nutrient accumulation (Ritchie and
McHenry 1990; Craft and Richardson 1993;
Reddy, et al. 1993). Atmospheric deposition
of 137Cs first began in 1954, with peak fall-
out occurring in 1964 (Ritchie and McHenry
1990). These two dates are most frequently
used to measure sedimentation rates (Ritchie
and McHenry 1990). However, the 1954 sedi-
ment horizon is sometimes difficult to discern
because of the low 137Cs activity (compared to
the 1964 peak), bioturbation and diffusion of
27
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137Cs (Ritchie and McHenry 1990). For this
reason, the 1964 peak, frequently is used as
the marker layer in wetland studies (Craft and
Richardson 1993a,b; Reddy etal. 1993). Thus,
the maximum 137Cs level in the soil profile
corresponds to the time of maximum deposi-
tion of 137Cs fallout, approximately 1964.
CALCULATING SEDIMENT
AND NUTRIENT
ACCUMULATION RATES
Nutrient (C, N, P) accumulation rates were
calculated as follow.
Nutrient Accumulation Rate = BD * NC
where:
BD = mean bulk density above the
137Cs peak
NC = mean nutrient concentration
above the 137Cs peak
Sediment accumulation rates are calculated
similarly by subtracting the dry mass of or-
ganic matter from the total dry mass of the
soil based on the assumption that soil organic
matter is 50 percent organic C.
POREWATER AND WATER
EXTRACTABLE NUTRIENTS
Movement of nutrients and other contami-
nants between the soil and water column
depends on their concentration in the soil
pore water (interstitial water) and in the wa-
ter column. Water content in wetland soils
may vary. Water content of mineral soils is
typically in the range of 30-50%, while or-
ganic soils may have up to 95% water con-
tent. Recently deposited flocculent sediments
may have water content of up to 99%. Some
of this water is held by soil particles through
capillary forces and trapped in the crystal
lattice of minerals. A major portion is present
as free water. This interstitial water, called
porewater, fills the space between soil par-
ticles. Sampling porewater is laborious and
requires special equipment, and may require
some experience and knowledge of the chem-
ical interactions in soils. For certain chemical
parameters, such as dissolved P, iron, manga-
nese, and sulfides, it is critical that porewater
extractions are performed under oxygen-free
conditions. Common methods used to extract
porewater include squeezing and centrifuga-
tion. Sulfide analysis must be performed on
unacidified samples. In situ porewater sam-
pling includes the use of porewater sippers or
porewater equilibrators commonly known as
"peepers" (Hesslein 1972). These methods
may not be suitable for routine sampling of
porewater due to disturbance and duration
of deployment. If the goal of sampling is to
determine the dissolved concentration of
metals, ammonium N, and P, samples must
be preserved under acid conditions (pH <2).
Analysis for these dissolved constituents can
be performed using standard methods as de-
scribed for water quality indicators.
Under certain conditions, it may not be
possible to extract porewater constituents,
especially during low water-table conditions.
Under these conditions, soil samples can be
extracted with distilled water at a soil (dry
weight) to water ratio of 1:4, after equilibra-
tion for a period of one hour on a shaker table.
Soil suspensions are centrifuged and filtered
through 0.45 um filter. The filtrates are acidi-
fied to pH <2 and stored at 4°C for analysis.
28
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EXTRACTABLE PLANT
AVAILABLE NUTRIENTS
AND METALS
(USING STANDARD SOIL
TEST PROCEDURES RELEVANT
TO THE REGION)
Selected methods used to determine P and
metals are described below. Many soil-test-
ing laboratories perform this analysis on a
routine basis. The following extractions can
be performed on air-dried soil that has been
ground to pass a 2 mm sieve.
Mehlich-I: The Mehlich-I extraction solution
consists of 0.05 MHC1 and 0.0125 MH2SO4.
It is typically used in the Southeast and Mid-
Atlantic regions on mineral soils with pH of
<7.0. The extractant consists of dilute con-
centrations of strong acids. Many of the plant
nutrients such as P, K, Ca, Mg, Fe, Zn, and
Cu extracted with Mehlich-I methods have
been calibrated for the production of crops
in agricultural ecosystems. This solvent ex-
tracts some of Fe and Al-bound P, and some
Ca-bound P. Soil (dry) to extraction ratio is
set at 1:4 (5 g of soil plus 20 ml of extract-
ant solution), for mineral soils. Wider ratios
should be used for highly organic soils. The
soil solutions are equilibrated for period of 5
minutes on a shaker table and filtered through
Whatman No. 42 filter. Filtered solutions are
analyzed for P and other nutrients using stan-
dard methods.
If the filtrate for a sample contains 10 mg
P/L, the soil contains 40 mg P/kg, as shown
below:
Mehlich-I-P = [(10 mg P/L) (0.020
L)]/0.005 kg soil = 40 mg
P/kg of soil
Mehlich-III: The Mehlich-III method was
developed to evaluate levels of available nu-
trients in soils of southeastern United States
(Mehlich 1984). The Mehlich-III extraction
solution consists of 0.2 M CH3COOH and
0.25 MNH4F, 0.013 MHNO3, and 0.001 M
EDTA. Inclusion of ammonium fluoride pro-
vides a better estimation of P availability in
near-neutral and alkaline soils than acid solu-
tions used in Mehlich-I method. The chelat-
ing agent EDTA aided in the extraction of
available metals. Many of the plant nutrients
(such as P, K, Ca, Mg, Fe, Zn, and Cu) ex-
tracted with Mehlich-III methods have been
calibrated for the production of crops in ag-
ricultural ecosystems. Soil (dry) to extraction
ratio is set at 1:8 (2.5 g of soil plus 20 ml of
extractant solution), for mineral soils, while
wider ratios should be used for highly organic
soils. Soil solutions are equilibrated for peri-
od of 5 minutes on a shaker table and filtered
through Whatman No. 42 filter. Filtered solu-
tions are analyzed for P and other nutrients
using standard methods.
Bray P-l: This method has been widely
used as an index of available P in soils. The
combination of dilute concentrations of a
strong acid, HC1 (0.025 M), with ammonium
fluoride (NFLF at 0.03 M} is designed to eas-
ily remove acid-extractable soluble P forms
such as Ca-bound P, and a portion of Fe and
Al-bound P. The NH4F dissolves Fe and Al-
bound P by forming complex ions with these
metal ions in acid solutions. This method has
been successfully used on acid soils. Soil
(dry) to extraction ratio is set at 1:7 for miner-
al soils, while wider ratios should be used for
highly organic soils. Soil solutions are equil-
ibrated for period of 5 minutes on a shaker
table and filtered through Whatman No. 42
filter. Filtered solutions are analyzed for P
and other nutrients using standard methods.
29
-------
Bicarbonate Extractable P: This method
is suitable for calcareous soils. Soil P is ex-
tracted from the soil with 0.5 MNaHCO3, at
nearly a constant pH of 8.5. In calcareous, al-
kaline, or neutral soils, containing Ca-bound
P, this extractant decreases the concentration
of Ca in solution by causing precipitation of
Ca as CaCO3 ; and as result P concentration
in soil solution increases. Soil (dry) to extrac-
tion ratio is set at 1:20 for mineral soils and
1:100 for highly organic soils. Soil solutions
are equilibrated for period of 30 minutes on
a shaker table and filtered through Whatman
No. 42 filter. Filtered solutions are analyzed
for P using standard methods.
materials in biological oxygen demand (BOD)
bottles for a 24-hour period (WBL 2001). A
known amount of wet soil (5-10 g) is weighed
into a 250 mL BOD bottle containing a stir
bar. The bottle is filled with deionized dis-
tilled water under continuous stirring on a
magnetic stirrer. The contents are stirred for a
period of 15 minutes and the DO is measured.
The bottle is sealed with a glass stopper and
placed in the dark at 25°C for a period of 24
hours. At the end of incubation, DO is mea-
sured under continuous stirring. If the DO
levels decrease by more than 50%, analysis
should be repeated with smaller sample size.
Soil oxygen demand is calculated as follows:
OXALATE EXTRACTABLE
ALUMINUM AND IRON
A number of researchers have shown a strong
relationship between the P retention capacity
of soils and the ammonium oxalate extract-
able Fe and Al content of the soils (Reddy,
et al. 1998b). This extractant dissolves poorly
crystalline and amorphous forms of Fe and
Al, which are primarily found to retain P in
acid mineral soils. Thus, measurement of ox-
alate extractable Fe and Al provides an index
of P retention capacity of soils. Air dried soils
are extracted with 0.1 Moxalic acid + 0.175 M
ammonium oxalate (pH = 3.5) at a soil to ex-
tractant ratio of 1:50. After 4 hour equilibra-
tion on a mechanical shaker table, soil sus-
pensions are centrifuged and filtered through
0.45 um filter. Filtered solutions are analyzed
for Fe, Al, and P using standard methods.
SOIL OXYGEN DEMAND
Soil oxygen demand (SOD) can be mea-
sured using an O2 electrode and incubation of
SOD (mg/kg-hour) = [{[DO, mg/L]t=0-
[DO, mg/L]^24hour j y L] / dry weight of soil, kg
Where: DO = dissolved oxygen;
V = volume of water in L, liters
LIGNIN AND CELLULOSE
CONTENT
(LIGNO-CELLULOSE INDEX)
Decomposition rate is significantly affected
by chemical composition of detrital tissue
and soil organic matter. Lignin and cellulose
comprise major components of organic mat-
ter, dictating the rate of decomposition and
substrate quality. Cellulose is more easily
biodegraded than lignin. The ligno-cellulose
index [LCI = (lignin/lignin + cellulose)] has
been used to characterize substrates for their
decomposability. The ligno-cellulose index
can be calculated from these measurements
and used to characterize the decomposability
of sampled soil material. For example, dur-
ing decomposition, the LCI for cattails was
shown to increase from 0.2 to 0.8, with low
30
-------
values observed in easily decomposable live
tissue and high values observed in recalci-
trant litter incorporated into soil organic mat-
ter (DeBusk and Reddy 1998).
DETRITAL/LITTER
DECOMPOSITION
(LITTER BAG MEASUREMENTS)
Decomposition of detrital/litter material
in wetlands results in the release of nutrients
into the water column. Ecologists have used
simple litter-bag methods to determine the
decomposition rate of plant tissue (Aber and
Melillo 1980; Wieder and Lang 1982). Typi-
cally detrital/litter material is chopped into
(approximately 2-5 cm in length), and known
amount of fresh plant tissue is placed in litter
bags (approximately 15 x 30 cm2) constructed
of fiber glass screening (1 mm mesh). The
edges of litter bags are stapled at 5 cm inter-
val, to provide relatively large openings along
the periphery for macroinvertebrate access.
The bags, with detrital tissue, are placed on
the soil surface or in the detrital layers. The
bags are randomly removed at predetermined
time intervals, carefully rinsed to remove ex-
ternal debris, and the detrital tissue is dried at
70°C for a period 72 hours, and dry weights
are recorded. Loss in dry weight of the litter
provides an indication of the rate of decom-
position.
COTTON-STRIP ASSAY
Soil organic matter decomposition is an
important process for the autochthonous mo-
bilization of nutrients for plant growth and
controlling organic matter accumulation. The
decomposition of cellulose strips has been
used extensively in the literature as a surro-
gate for plant organic matter decomposition,
providing a method for normalizing substrate
quality between sites (French 1988; Harrison,
et al. 1988). The cotton strip technique for
quantifying cellulose decomposition is based
on the loss of tensile strength of cellulose
fibers, referred to as cotton tensile strength
loss (CTSL), of a standardized cotton fabric
comprising 97% holocellulose (Latter and
Howson 1977; Latter and Harrison 1988).
The cotton strip technique evaluates decom-
position by measuring loss of tensile strength
of the cotton fibers making up the strips.
Measurements of tensile strength loss rates
have been undertaken and proven success-
ful in a wide range of wetlands (Maltby 1987;
Newman, et al. 2001) and non-wetland envi-
ronments (Harrison et al. 1988) throughout
the world. At each location, replicate cotton
strips of the standard material (Shirley Insti-
tute Test Fabric, Didsbury, England) and size
(12 x 30 cm2 or longer depending on objec-
tives) are inserted vertically into the soil sub-
strate with the aid of a sharpshooter shovel
as described by Maltby (1987). Control strips
are inserted and removed immediately. The
remaining strips are exposed for usually a
10-14 day period, depending on the rate of
decomposition. After retrieval, all strips are
immediately washed in freshwater to remove
soil and debris and washed again in deionized
water. Samples are dried at room temperature
and stored in plastic bags until analysis. The
strips are cut into horizontal segments 3 cm wide
and reduced by fraying to 2 cm segments that
result in test units corresponding to depths
of 0-2, 3-5, 6-8 cm, etc. Loss of tensile
strength is measured from each 2 cm x 12 cm
segment. Tensile strength is measured with a
tensometer (Monsanto Type-W or equivalent)
equipped with 7.5 cm wide jaws adjusted to
3 cm spacing. All measurements are carried
31
-------
out at 18-22°C and 100% relative humidity
obtained by soaking the strips in deionized
water. Individual losses in tensile strength
are calculated relative to the field controls
obtained for each site. These data are used
to calculate percentage mean loss of tensile
strength for each level in the profile. This pro-
cedure is sensitive to differences in soil fertil-
ity (Maltby 1985, 1987; Mendelssohn, et al.
1999).
MICROBIAL ACTIVITIES
Microbial parameters (listed in Table 1,
Level II) are now routinely measured by
researchers. Examples of some of these pa-
rameters include: (i) extracellular enzyme
activity (Wright and Reddy 2001a); (i) mi-
crobial biomass C, N, and P (Ivanoff, et al.
1996; White and Reddy 2000); (ii) microbial
respiration and methanogenesis (Wright and
Reddy 2001b); (iii) potentially mineralizable
N (Wright and Reddy 2000); (iv) potential-
ly mineralizable P; and, (v) denitrification
enzyme activity (White and Reddy 1999).
Details about methodologies can be obtained
from the references cited.
MINIMUM MONITORING
REQUIREMENTS
A Nutrient related data on water and soil
L \ quality indicators in wetlands is lim-
ited. To date, much of the data collection is
at the experimental scale for site-specific
conditions. Assessing wetland eutrophica-
tion requires systematic data collection at a
large spatial scale, using comparable tech-
niques. Recognize that under most condi-
tions adequate resources might not be avail-
able to obtain detailed data even for Level I
indicators. However, wetland nutrient assess-
ment requires minimum data collection and
evaluation of soil and water quality indica-
tors. This document presents a simple, sys-
tematic approach to minimal data collection.
Level I indicators required for minimum data
needs are listed in Table 1.
The location of suitable field sites should be
coordinated through local academic or gov-
ernmental agencies to determine appropri-
ate reference sites for sampling (see Module
#4—Study Design for Monitoring Wetlands).
Criteria for reference site selection should be
based on areas of least cultural impact within
a particular region as determined by the lo-
cal knowledge source. Once identified, sites
should be characterized using a standard clas-
sification scheme (Cowardin, et al. 1979) and
by community characteristics. Latitude and
longitude of the site should also be collected
for use in relocation and cross-referencing
with other geographic information system
(GIS) sites.
Sampling at selected sites should consist
of three composite samples collected from
the water column, detritus and soil. Water
samples, when available, should be collected
at a mid-water depth, filtered, homogenized
(with other replicates) then stored on ice or
preserved until analysis. Intact soil should be
collected to a depth of 10 cm below the lit-
ter/soil interface. Litter from these cores, as
defined by easily distinguishable plant frag-
ments lying on the surface of the core, should
be collected, air dried, and then combined
with other detritus samples from the site. The
remaining upper 10 cm of soil from each core
should be air dried, then combined with site
replicates.
32
-------
Laboratory analysis of water column com-
posite samples should include total nitrogen
and total phosphorus. Soil and detritus com-
posite samples should be air-dried at 25 to
30°C, ground, and homogenized. Aliquots
should be analyzed for organic matter con-
tent, pH (to be determined on ambient wet
sample), total nitrogen, total phosphorus, ex-
tractable ammonium N (to be determined on
ambient wet sample), and extractable phos-
phorus, iron, aluminum, calcium, and mag-
nesium. Moisture content of air-dried detrital
matter and soils should be determined after
over drying samples at 70°C for 2-3 days or
until constant weights are recorded. All nu-
trient concentrations determined on air-dried
samples should be normalized on an over-
dried basis.
CASE STUDY:
THE EVERGLADES
rhe Everglades is one of the most unique
subtropical wetland ecosystems in the
world. It evolved biologically from an organic
Minimum Data - Level I
Water column:
& Water depth
& Total Nitrogen
& Total Phosphorus
Litter/Detritus:
& Total carbon,
& Total nitrogen
& Total phosphorus
Soil:
& Bulk density
& Organic matter content,
& Total carbon,
& Total nitrogen,
& Total phosphorus,
ts Extractable nitrogen (2M KCI),
& Extractable phosphorus (Mehlich
ts Extractable Fe, Al, Ca, Mg, and K
-1 and 3)
matter accumulation in a low-nutrient envi-
ronment sitting within a limestone depression
(Davis 1943). Historically, the major source
of nutrients to the Everglades has come from
atmospheric deposition, with minimum sec-
ondary nutrient inputs through infrequent
sheet flooding in the northern Everglades
from Lake Okeechobee. Nutrient limitation,
hydrology, and fire are several key factors in
the establishment of the endemic Everglades
flora, which has adapted to the low nutrient
environment (Davis 1991). Nutrient loading
to Water Conservation Areas (WCAs) of the
northern Everglades has not only altered the
vegetational communities, but also increased
nutrient accumulation (Davis 1991; DeBusk,
etal. 1994; Newman, etal. 1997). Although the
effects of nutrient loading and altered hydrol-
ogy on changes in plant communities are
clearly evident, very limited information is
available on the influence of these factors on
biogeochemical processes regulating nutrient
availability and cycling in impacted and un-
impacted areas. Nutrient accumulation rates
of 0.11-1.14 g nr2 yr1 potassium and 5.4-24.3
g m~2 yr"1 nitrogen have been reported for the
Everglades (Craft and Richardson 1993a,b;
Reddy, et al. 1993). The
highest accumulation
rates were noted in areas
closest to the source of
nutrient inputs, and the
lowest accumulation rates
occurred in areas furthest
from the input points.
A brief summary of the
results of various bio-
geochemical indicators
measured along a nutri-
ent enrichment gradient
in WCA-2a of the Ever-
glades are listed in Table
2 (Reddy, et al. 1999).
33
-------
TABLE 2A: BACKGROUND CONCENTRATIONS (STANDARD ERROR) SELECTED
BIOGEOCHEMICAL INDICATORS IN FLOC/DETRITAL COMPONENT OF VARIOUS
HYDROLOGIC UNITS OF THE EVERGLADES. THE PMN AND PMP REFER TO
POTENTIALLY MINERALIZABLE N AND P, RESPECTIVELY (WRIGHT, ET AL. 2OO2)
PARAMETER UNITS WCA-1 (SE) WCA -2A WCA-3A TS
FLOC/ DE TRITU S
Los s on Ignition
Extractable C
Extractable NH4-N
Labile P
To tal P
To tal inorganic P
To tal N
To tal C
MicrobklBiomass C
MicrobklBiomass N
Microbkl Biomass P
PMN
PMP
Soil O2 Demand
CH4 Production
Aerobic CO2Prod.
Anaerobic CO2Prod.
%
g C kg-1
mg N kg'1
mg P kg*1
mg P kg'1
mg P kg-1
gkg'1
gkg'1
g C kg-1
mg N kg'1
mg P kg-1
mg N kg'1 d'1
mg P kg-1 d'1
mg kg-1 hr'1
mg C kg-1 d'1
mg kg-1!} -1
mgkg-1!!-1
83
7
4
227
51
35
413
24
1346
305
.
53
36
(3)
(1)
(4)
(21)
(6)
(1)
(9)
(14)
85
(65)
(1)
(11)
50
5
181
1
315
157
18
261
14
1294
117
88
5
118
36
34
(3)
(0)
(8)
(1)
(22)
(11)
(1)
(11)
(2)
(202)
(17)
(10)
(2)
(?)
(4)
(15)
91
12
379
6
540
174
45
448
30
2879
87
212
24
89
564
(0)
(1)
(40)
(1)
(16)
(15)
(0)
(3)
(5)
(250)
(22)
(25)
(8)
(12)
(53)
39
7
177
1
156
58
17
250
10
1171
53
49
2
26
181
(3)
(1)
(8)
(0)
(9)
(4)
(1)
(9)
(3)
(334)
(25)
(7)
(0)
(3)
(23)
Observations made in the Everglades
wetlands are applicable to nutrient limited
systems. Although, the indicators presented
in this report will be same for any wetland,
their relative rates and sensitivity will be dif-
ferent, depending on type of wetland and nu-
trient condition. Thus, the application of the
Everglades results to other wetlands should be
viewed as starting point and caution should
be exercised in selecting these indicators.
Several biogeochemical parameters and as-
sociated processes are affected by P loading
to P limited areas of the Everglades. A sum-
mary of the results for the WCA-2a sites re-
ferred as impacted and reference conditions
are describe below (Reddy et al. 1999). Total
and bicarbonate extractable P were higher in
the detrital layer and soils of the impacted
site than the unimpacted site. The C:P ratio
of detritus and soils decreases by over 50%
34
-------
TABLE 2B: BACKGROUND CONCENTRATIONS (STANDARD ERROR) SELECTED
BIOGEOCHEMICAL INDICATORS IN SURFACE SOIL (O-3 CM) COMPONENT OF
VARIOUS HYDROLOGIC UNITS OF THE EVERGLADES. THE PMN AND PMP REFER
TO POTENTIALLY MINERALIZABLE N AND P, RESPECTIVELY. (WRIGHT, ETAL. 2OO2)
PARAMETER UNITS WCA-1 (SE) WCA-2A WCA-3A TS
SOILf 0-3 CM)
Loss on Ignition
Extractable C
Extractable NH4-N
Labile P
TotalP
To tal inorganic P
TotalN
To talC
Microbial Biomass C
Microbial Biomass N
Microbial Biomass P
PMN
PMP
Soil O2 Demand
C H 4 P rod uc tion
Aerobic CO2Prod.
Anaerobic CO2Prod.
%
gCkg-1
mg N kg"1
mg P kg-1
mg P kg"1
mg P kg-1
g N kg-1
gCkg-1
KG kg-1
mg N kg-1
mg P kg-1
mg N kg-1 d"1
mg P kg-1 d-1
mg kg-1 hr1
mg C kg-1 d-1
mg kg-1!! -1
mg kg-1!! -1
91
3
116
2
223
45
36
459
3
194
77
27
2
14
13
10
(0)
(0)
(5)
(1)
(10)
(3)
(1)
(7)
(1)
(38)
(7)
(3)
00
(4)
(5)
(2)
72
3
112
1
358
153
29
360
3
294
76
35
4
21
18
13
(3)
(0)
(14)
(0)
(27)
(16)
(2)
(13)
(0)
(46)
(8)
(7)
(1)
(3)
(6)
(3)
90
5
109
8
360
103
41
463
3
597
24
24
4
33
105
(1)
(1)
(6)
(3)
(10)
(7)
(1)
(3)
(1)
(97)
(6)
(4)
(0)
(3)
(13)
20
2
29
1
120
50
10
193
2
317
20
9
2
15
21
(2)
(0)
(3)
(0)
(11)
(4)
(1)
(8)
(0)
(85)
(3)
(2)
(0)
(5)
(2)
as a consequence of P inputs into the system.
Clearly, P loading has enriched the soil forms
of P. Microbial biomass (MB) C, N and P were
also higher in the detrital layers and surface
soils (0-10 cm depth) in P enriched areas.
Phosphorus to microbial biomass, expressed
as the ratio of MBP/Ptotal in the detrital layer,
decreased from 27% to 16% in the impacted
area. A similar 50% decrease in MBP/P ,
total
was observed in the 0-10 cm soil interval. In-
creased MB has led, in turn, to higher rates
of microbially mediated processes that regu-
late the biogeochemical cycling of C, N and P.
Breakdown of organic matter had increased,
evidenced by greater microbial respiration
rates and higher activity of some extracel-
lular enzymes. The net mineralization rates
or releases of inorganic N and P were higher
in soils and detrital layers at the elevated P
site, increasing nutrient availability to higher
plants. Nitrogen fixation, nitrification and
denitrification rates also increased in the im-
35
-------
Detrital/Soil Impact Index Relative sensitivity to
Biogeo chemical Indicator/Process
layer log [TS/RS] phosphorus loading
APA
C/P ratio
SIPM
(SIPM/MBP)
(MBP/TP)
APA
(MBP/TP)
Ash content
Metabolic quotient
Arylsulfatase
Ash content
To tal C and N
Metabolic quotient
C/N ratio
Extra ctable NH4-N
SIPM
PMP
General aerobes
Protease
Protease
Aerobicmicrobial respiration
MBC
MEN
Extra ctable NH4-N
MBN/N total
Nitrification
Arylsulfatase
Phenol oxidase
To tal P
B-D glucosidase
MBC
Anaerobic microbial respiration
PMN, SINM
Nitrification
N2 fixation
DBA
SINM/MBN
MBP
To tal P
B-D glucosidase
N2- fixation
SINM
Bicarbonate extractable P
PMP
D
D,S
S
D,S
S
S
D
D
D
D
S
D,S
S
D,S
S
D
S
S
S
D
D,S
S
D,S
D
D,S
D
S
D,S
S
S
D
D,S
D,S
D
D
D
D,S
D
D
D
D
D
D,S
D
-1.0 to -0.50
-0.50 to -0.25
-0.25 to -0.10
-0.10 to 0.10
0.10 to 0.25
0.25 to 0.50
0.50 to 1.0
High
Medium
Low
Negligible
Low
Medium
High
36
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TABLE 3: IMPACT INDICES AND RELATIVE SENSITIVITY OF VARIOUS
BIOGEOCHEMICAL PROCESSES/INDICATORS MEASURED IN DETRITAL AND
SOIL LAYERS AT IMPACTED AND REFERENCE SITES IN THE EVERGLADES.
D = DETRITAL LAYER AND S = O-1O CM SOIL (REDDY, ET AL. 1999).
Detrital/Soil Impact Index Relative sensitivity to
layer log [TS/RS] phosphorus loading
Anaerobes*
Acetate producers
H2 consumers
CO2 consumers
Methanogens
Sulfate reducers
s
s
s
s
s
>1.0
Ve ry H igh
^Anaerobes were notmeasured in the detrital layer.
APA = alkaline phosphatase activity; C/ P — c arbon to phospho rus mass r atio; SIPM = substrate induced organic P mineralization;
MBP = microbialbiomass P; PMP = potentially mineralizable P; MBC = m krobialbiomass C; MBN = m krobial bio mass N; PMN
= potentially mineralizable N; SINM = substrate induced organic N mineralization; and DEA = denitrification en zyme activity
pacted area. Overall, P loading increased the
size of the microbial pool and organic matter
turnover rate, which led to a greater release
of inorganic N and P, further driving eutro-
phication.
These indicators were useful in evaluating
impacts of nutrient loading on the ecosystem's
health and depended on an indicator's natural
variability within the system. Uncertainty in
evaluation was caused by spatial variability,
as well as the dynamic nature of many of the
processes, complicating the extrapolation of
laboratory results to field conditions. Temporal
variation in these indicators introduced addi-
tional uncertainty in the interpretation of re-
sults. In an ecosystem with distinct gradients,
impacts were described by using the rela-
tionship developed between biogeochemical
processes and associated easily measurable
parameters. However, these relationships
needed to be based on data collected on sites
having a wide range of physical, chemical and
biological properties, and loading impacts.
For comparison purposes, a simple impact
index was calculated in the Everglades for
each process or parameter measured:
Impact index = log [TS/RS]
Where [TS] is the rate or concentration of
a parameter measured at an impacted or test
site; and [RS] is the rate or concentration of
a parameter measured at a reference site. The
log [TS/RS] provided an index value of zero,
which indicated no change, a negative value
indicated a decrease and a positive value
indicated an increase. For example, a value
of "1" represented a ten-fold change in con-
centration of a parameter or rate of a process
at the impacted site, relative to the reference
site. This approach allowed the ranking of
the parameters or processes most affected
by nutrient loading/disturbance. Impact in-
dices were viewed in the context of spatial
variability within impacted and unimpacted
sites. Field replicates may vary in relatively
small areas (
-------
the detrital layer ranged from 27% to 100%,
whereas variability in microbial respiration
rates was <20% (Wright and Reddy 2001,
2002). For measurements of indicator param-
eters collected at WCA-2a sites, variability in
experimental techniques was <10% for both
in situ and laboratory conditions.
The calculated impact index values, shown
in Table 3, aided in normalizing rates of bio-
geochemical processes and concentrations of
various parameters into a common format.
Data presented in this paper could help when
developing a qualitative ranking of processes
and parameters based on their sensitivity to
P loading. Because of the variability of un-
certainty, the index values were grouped into
four broad categories of positive and nega-
tive impacts. We assumed that impact index
values in the range of-0.1 to 0.1 were within
the experimental variability of many of the
processes and parameters measured. For ex-
ample, the variability of the impact index
for total P was <0.2 for the detrital layer and
<0.1 for the surface soil of the reference site,
respectively. However, for some parameters
such as microbial numbers (usually reported
on log-scale), the variability was much higher.
Biogeochemical processes and parameters
measured on the detrital layer were more sen-
sitive to nutrient loading than those measured
on surface soil. The detrital component proba-
bly represented most recent (non-steady state)
impacts, whereas the surface soil likely rep-
resented longer-term conditions that may not
have responded to impact yet. Microbial com-
munities associated with detrital layer were
in direct contact with water column nutrients,
and should respond rapidly to changes in
water chemistry. However, impact index val-
ues obtained on soils may be more reliable
because they represented long-term steady
state conditions.
These results provided a simple strategy for
integrated evaluation of P impacts on wet-
lands, using soil biogeochemical processes
and parameters as potential indicators. Al-
though biogeochemical processes may be
sensitive and reliable indicators of wetland
integrity, their measurements were time-con-
suming and expensive. We have found, how-
ever, that concentrations of certain chemical
substrates, intermediates, and end products
functioned as surrogates for biogeochemical
processes (Reddy and D'Angelo 1996; Reddy,
et al. 1998; D'Angelo and Reddy 1998; De-
Busk and Reddy, 1998). Furthermore, rela-
tionships between indicators and processes
provided reliable estimates of ecosystem
health. The simple strategies presented in
this paper provided a tool to normalize the
process level information into a common
format for possible integration into predic-
tive models. Evaluating impacts required a
reliable database for several reference sites
(background level of impacts) in various geo-
graphical regions as classified by wetland veg-
etation, soils, and hydrology. The strategies
presented could be used in different types of
wetlands to assess restoration and remediation
efforts, and potentially be used as a screen-
ing tool to choose how best to utilize limited
restoration resources.
38
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