Citation
Spatial and temporal distribution of mercury and other metals in Florida Everglades and Savannas marsh soils

Material Information

Title:
Spatial and temporal distribution of mercury and other metals in Florida Everglades and Savannas marsh soils
Creator:
Rood, Brian Eugene, 1963-
Publication Date:
Language:
English
Physical Description:
viii, 180 leaves : ill. ; 29 cm.

Subjects

Subjects / Keywords:
The Everglades ( local )
Mercury ( jstor )
Sediments ( jstor )
Atmospherics ( jstor )
Genre:
bibliography ( marcgt )
theses ( marcgt )
non-fiction ( marcgt )

Notes

Thesis:
Thesis (Ph. D.)--University of Florida, 1993.
Bibliography:
Includes bibliographical references (leaves 124-142).
General Note:
Typescript.
General Note:
Vita.
Statement of Responsibility:
by Brian Eugene Rood.

Record Information

Source Institution:
University of Florida
Holding Location:
University of Florida
Rights Management:
Copyright [name of dissertation author]. Permission granted to the University of Florida to digitize, archive and distribute this item for non-profit research and educational purposes. Any reuse of this item in excess of fair use or other copyright exemptions requires permission of the copyright holder.
Resource Identifier:
030440173 ( ALEPH )
AKC8876 ( NOTIS )
31274879 ( OCLC )

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Full Text











SPATIAL AND TEMPORAL DISTRIBUTION OF MERCURY AND OTHER
METALS IN FLORIDA EVERGLADES AND SAVANNAS MARSH SOILS

















By

BRIAN EUGENE ROOD


A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL
OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT
OF THE REQUIREMENTS FOR THE DEGREE OF
DOCTOR OF PHILOSOPHY

UNIVERSITY OF FLORIDA


1993































Copyright 1993

by

BRIAN EUGENE ROOD































This dissertation is dedicated to my parents, F. Eugene and Roberta Rood, the best

teachers I've known. I also dedicate this dissertation to Professors Edward S. Deevey, Jr.

(deceased) and Peter H. Rich for prompting me to recognize the power of the imagination.














ACKNOWLEDGMENTS


I wish to thank Dr. Joseph Delfino, for his supervision of my research, and my

committee members, Drs. Ronnie Best, Emmett Bolch, Donald Graetz, and Frank Nordlie

for their critical review of this dissertation. Dr. Claire Schelske kindly reviewed my

dissertation and attended my oral defense. I gratefully acknowledge laboratory assistance

by William Beddow, Candace Biggerstaff, Celia Earle, Becky Fierle, Ingrid Forbes,

Lizanne Garcia, Manuel Llahues, Kathleen Newell, Margaret Olson, Brandon Selle,

Marcia Sommer, and Melissa Voss. Special thanks go to Richard Pfeuffer, Liberta Scotto,

and Bob Przekop for their assistance with planning and implementation of field sampling,

and to Curtis Watkins, Dr. Thomas Atkeson, Thomas Swihart (Florida Department of

Environmental Protection) and Larry Fink (South Florida Water Management District),

who served as project officers for the funding agencies. This research was funded by

grants from the Florida Department of Environmental Protection, South Florida Water

Management District, and the United States Geological Survey. I am greatly indebted to

my friend, Dr. Johan F. Gottgens, for his assistance, generosity, and candor, throughout

my doctoral studies, and to Brian Cutchens and Lola Wilcox for their treasured friendship.














TABLE OF CONTENTS



ACKNOWLEDGMENTS .......

A B ST R A C T ... .. .. .. ........... ....... ... ... ... .

CHAPTER 1
INTRODUCTION ...........
Background ..................
Present Study ...................

CHAPTER 2
REVIEW OF LITERATURE .............
O verview ............................
Human-Related Activities ............................
M ercury Issues in Florida ....................... .
Available Technology for Mercury Research ...............
Global Mercury Cycle ................
Global and Regional Interactions ...........
Mercury in the Atmosphere .............
M ercury in W ater ..................................
Mercury in Sediment ................................
M ercury in B iota ...................................
Environmental Factors and Bioaccumulation .......
Mercury Transformations in Aquatic Systems ....
Bioaccumulation in Fish ...........


Identification and Assessment
Paleolimnological Studies ........
Sum m ary ...................

CHAPTER 3
MATERIALS AND METHODS ........
Site Selection ................
Field Sampling ...............
Total M ercury ................
Percent Solids/Bulk Density ......
Radionuclide Analysis ..........


of Mercury Contamination









C arbon ................................ ......... ....... 49
Total Carbon ....................... ...... ..... 49
Inorganic and Organic Carbon ........................ 50
Additional Trace M etals .............................. .... 50

CHAPTER 4
RESULTS AND DISCUSSION ................................... 53
W ater Quality ......... ............. ....... ........ 53
Sediment Geochronology ............ .. ............. 56
Sediment Dating Acceptance Criteria ................ 58
Error Analysis of Sediment Dating ....................... 86
Sediment Mercury Concentrations ................. ........ 94
Comparison of Recent and Historic Mercury Concentrations .... 94
Post-Depositional Mobility of Mercury .................. 99
Error Analysis of Mercury Determinations ... ... ...... 102
Spatial Distribution of Mercury in the Everglades ......... .. 102
Relationships Between Mercury Concentration and Selected Water
and Sediment Parameters ... ................. 106
Supplementary Sediment Metals Concentrations ................... 108

CHAPTER 5
SUMMARY, CONCLUSIONS, AND RECOMMENDATIONS ............. 120
Sum m ary .. .. .. .. .. .. .. 120
Conclusions ....................................... 121
Recom m endations ........................................ 122

REFERENCE LIST ............................................ 124

APPENDIX
FLORIDA WETLAND SOIL CHEMISTRY DATABASE ................. 142

BIOGRAPHICAL SKETCH ...................................... 180














Abstract of Dissertation Presented to the Graduate School
of the University of Florida in Partial Fulfillment of the
Requirements for the Degree of Doctor of Philosophy

SPATIAL AND TEMPORAL DISTRIBUTION OF MERCURY AND OTHER
METALS IN FLORIDA EVERGLADES AND SAVANNAS MARSH SOILS

By

Brian Eugene Rood

December 1993


Chairperson: Joseph J. Delfino
Major Department: Environmental Engineering Sciences

Elevated mercury concentrations were identified previously in freshwater fish in

the Everglades, Savannas State Reserve, and receiving waters of the Okefenokee Swamp.

The goals of this research were to 1) determine historic baseline concentrations of

mercury in Florida wetland soils, 2) determine post-development changes in sedimentary

mercury accumulation, and 3) identify the spatial distribution of mercury throughout the

Florida Everglades. Sixty soil cores were analyzed for total mercury. Selected cores

were analyzed for carbon and trace metals, and were chronologically analyzed after

radionuclide analysis for 2toPb and '"Cs.

The average mercury concentration in surface sediment (0-4 cm) of 121 ng g'

(n=51, 17-411 ng g-1) was 2.5 times (0.2-10.6, n=51) higher than corresponding deep

sediment (11-17 cm) concentrations. The largest increases were measured in Water









Conservation Areas 1 and 2 (3.7 times higher for both) of the Florida Everglades, while

Okefenokee Swamp sediment showed the smallest relative increase (1.4). Because

concentration data are vulnerable to temporal variations in bulk sediment accumulation

rate, the interpretive problem of co-variance was avoided by determining mercury

accumulation rates after radionuclide dating. Post-1985 mercury accumulation rates

averaged 53 Lg m2 y-i (23-141 gg m-' y"') corresponding to a 6.4 (1.6-19.1, n=18) times

rate increase since the year 1900. The largest rate increases occurred in WCA and

WCA2 cores (7.8 and 8.7 times higher, respectively), while the Savannas State Reserve

cores showed the smallest rate increase (3.4). Mercury accumulation rates increase

starting about 1940, due perhaps to mid-century alteration of the hydrologic structure of

the Everglades, and to increased regional agricultural and urban development. There is

presently insufficient information regarding regional inputs to quantify any direct causal

relationship between mercury accumulation rate increases and regional human activities.

However, apparent nonuniform accumulation of mercury in the Everglades hydrologic

basins, coupled with increased accumulation rates of other trace metals, indicate some

atmospheric contribution of mercury from regional anthropogenic activities. The findings

are similar to trends reported for lakes in Minnesota, Wisconsin, and Sweden. This

agreement is significant, perhaps indicating a global process that leads to similar

accumulation rates over widely varying geographic regions. This research provides the

first data on mercury accumulation in subtropical wetland systems and demonstrates the

feasibility of radiochemical dating of wetland cores.














CHAPTER 1
INTRODUCTION


Background


A statewide survey of mercury concentrations i sportfish was implemented after

preliminary indications of mercury contamination appeared in Florida freshwater fish

(Hand and Friedemann, 1990). The survey revealed mercury concentrations m fish in the

Everglades (Water Conservation Areas 1 and 2) and Savannas State Reserve that exceeded

acceptable levels for human consumption. Numerous lakes, rivers, and wetlands yielded

fish with mercury concentrations sufficient to warrant limited-consumption advisories.

This survey identified the magnitude of fish mercury contamination in the state.

However, the survey did not address issues regarding the origin, transport, and availability

of mercury in these habitats.

Concurrent studies of wildlife suggested that mercury is transported through the

food web of the Florida Everglades and that the viability of the endangered Florida

panther has been diminished due to mercury bioaccumulation (Roelke et al., 1991). The

risk to other animal populations from mercury biomagnification has been suggested and

the potential for perturbations of ecosystem structure and function have been examined

(Jurczyk, 1993).








2

Recent increases of mercury accumulation rate have been reported for north

temperate lake systems in Sweden. Wisconsin, and Minnesota (Meger, 1986; Wiener et

al., 1990; Lindqvist et al., 1991; Swain et al., 1992). In some cases, atmospheric

deposition could account for increased mercury accumulation rates in recent sediment

(Meger, 1986). Some studies have linked a 1.5% annual increase of atmospheric mercury

concentrations (1977-1990) (Slemr and Langer, 1992) to an estimated 2% increase in

mercury deposition rates in Wisconsin and Minnesota (Swain et al., 1992). These studies

suggested that mercury deposited on the surface and watershed of remote lake systems

originated from regional or global atmospheric sources (Swain et al., 1992).

It is estimated that about 95 percent of atmospheric mercury occurs in the gaseous

elemental form, with an atmospheric residence time of 0.7-2.0 years (Nater and Grigal,

1992). Approximately 5 percent of atmospheric mercury is associated with particulates

(Fitzgerald et al., 1991) which can readily be deposited as dryfall or scavenged from the

atmosphere during rain episodes (Fitzgerald, 1986). Anthropogenic emissions of

elemental mercury enter the global atmospheric cycle and may be distributed far from

their source; however, particulate phase mercury from emission sources may establish

regional concentration gradients in nearby soils (Nater and Grigal, 1992).

Urban emissions of mercury from industry (i.e. cement production), medical and

municipal waste incineration, fossil fuel combustion, in addition to agricultural emissions

(i.e. burning of crop material, volatilization of mercurial fungicides), may contribute to

atmospheric emissions and eventual deposition of mercury (Crockett and Kinnison, 1977;

Fukuzaki et al., 1986; Sengar et al., 1989; KBN Engineering and Applied Sciences, Inc.,

1992).








3

Perturbations of the natural hydroperiod may facilitate the release of historically

accumulated mercury because of changes in the physical properties of soils. Oxidation

and deep cracking of dried agricultural land may release both naturally and

anthropogenically derived mercury and facilitate its transport to wetland soils in runoff

(Del Debbio, 1991). Because flooded peat soils readily accumulate trace metals by

adsorption and sulfide precipitation they serve as a primary sink for mobilized mercury

(Lodenius et al., 1987, Norton et al., 1990).


Present Study


The Florida Everglades, Savannas State Reserve, and the receiving waters of the

Okefenokee Swamp (Suwannee and Santa Fe rivers) exhibited elevated fish mercury

concentrations (Hand and Friedemann, 1990). These aquatic systems are unique Florida

habitats, and there is concern that mercury contamination poses a serious ecological and

human health hazard. This study examines mercury abundance and distribution in

sediment from the Everglades, Savannas Marsh, and Okefenokee Swamp wetland systems

(Figure 1.1).

The Everglades is "perhaps the most recognized wetland in the world, its notonety

derived from the wealth of its biotic heritage as well as the magnitude of factors that

threaten its resources" (Gunderson and Loftus, 1993, p. 1). It is a dynamic subtropical

aquatic system (5600 km2), subject to hydrologic variability, fire, and human related

activities (Blake, 1980). The Everglades are considered oligotrophic, based on dominant

plant communities and ambient nutrient concentrations, and are characterized by peat soils



















Okefenokee


Savannas


Everglades


Figure 1.1. Geographic distribution of wetland study sites.








5

to the north and marl sediment to the south. Sawgrass (Cladium jamaicense) marshes

dominate large expanses of this system. The region is spotted with intermittent wet

prairies, tree islands and shallow ponds. During the past century, extensive draining of

this wetland for agriculture and diversion of water to coastal urban centers has altered its

natural hydroperiod. Regions south of Lake Okeechobee were drained for agriculture, and

canal systems were constructed to control water movement. Presently, some areas of the

remaining wetland are subject to prolonged dry periods while other locations are subject

to extended periods of inundation (SFWMD, 1992).

The Okefenokee Swamp, in southeastern Georgia and northern Florida, is the

second largest wetland in the United States (1750 km2). The flat, sandy watershed of the

Okefenokee Swamp is small (1200 km2) and siltation is negligible (Casagrande and

Erchull, 1976). As a result, precipitation serves as the predominant hydrologic input and

filling of the wetland basin is minimal. The Okefenokee consists of an "array of diverse

habitats" including lakes, wet prairies with floating peat mats (Sphagnum spp.), and

Taxodium spp. swamps that are integrated hydrologically to form one unit ecosystem.

This swamp has organic-rich soils underlain by a pure white quartz sand (Casagrande and

Erchull, 1976). The relatively pristine condition of the Okefenokee Swamp permits it to

"serve as a control for comparison with other ecosystems that continue to be heavily

influenced by human activities" (Rykiel, 1984, p. 374).

The Savannas State Reserve is a dynamic, linear wetland system (20 km x 2 km)

just west of the Indian River Ridge in Florida's St. Lucie and Martin counties. It is a

strip of marshlands, ponds, lakes, and islands, perched -4 m above mean sea level, and








6

characterized by rich inundated muck soils overlying relict sand dune on hardpan. The

marsh is dominated by broomsedge (Andropogon virginicus), water lily (Nymphaea

odorata), and spatterdock (Nuphar luteum) while the surrounding watershed is a pine

(Pinus elliottii) and saw palmetto (Serranoa reopens) habitat (Jurgens, 1981). This region

is considered to be "highly susceptible to damage by pollution or over-enrichment of its

water" (Davis, 1990, p. 4) due to its size and to encroaching development.

Previous studies have identified atmospheric mercury deposition as a primary

vector leading to mercury accumulation m aquatic systems (Meger, 1986). Recent

increases in mercury accumulation in aquatic systems have been attributed to global

(Swain et al., 1992) and regional (Sengar et al., 1989; Nater and Grigal, 1992) increases

of atmospheric mercury emissions, largely attributed to a variety of anthropogenic

activities (KBN Engineering and Applied Sciences, Inc., 1992). Further, elevated fish

mercury concentrations have been attributed to increased mercury inputs from human-

related activities (Bodaly et al., 1984; Hakanson et al., 1990a, 1990b). The following

study arose from concern that increased mercury inputs, of global or regional origin, were

causing elevated mercury concentrations in fish.

I hypothesize that the sediment record of these subtropical wetlands will concur

with previous indications, identified in other aquatic environments, of increased mercury

accumulation since the turn of the century. This hypothesis necessitates a characterization

of the feasibility of radiochemical dating in the study wetlands. In addition, identification

of spatial variations of mercury content throughout the Everglades is essential to

characterize the relative impact of regional activities on mercury accumulation in that

system.








7

This study of Florida wetland soils was initiated in 1991 to: 1) determine the

spatial distribution of mercury throughout the Everglades, Okefenokee Swamp, and

Savannas Marsh systems, 2) identify historic baseline concentrations of mercury in Florida

wetland soils, 3) identify post-development changes in sedimentary mercury accumulation,

4) identify mercury-organic associations in wetland soils, and, 5) provide information to

serve as a basis for informed planning and implementation of future research and

management activities. An evaluation of spatial and temporal changes in sedimentary

mercury is necessary to elucidate the factors governing mercury accumulation and

distribution in these wetland systems.














CHAPTER 2
REVIEW OF LITERATURE


Overview


Mercury is present in air, water, soil/sediment, and biota, and is unique among the

metals with its ability to exist in the gas, liquid, and solid phases (Clarkson et al, 1984;

Moore and Ramamoorthy, 1984). The abundance of mercury in the environment is

determined by the inputs supplied by both natural and anthropogenic processes

(Fitzgerald, 1986; Mitra, 1986). Natural processes, such as volcanism and degassing from

the land and ocean, supply a baseline mercury contribution to the atmosphere and to

water. That mercury is eventually transported to, and accumulated in, soil, sediment, and

biota. Mercury inputs to the environment may undergo numerous transformations that are

determined by physico-chemical interactions of mercury under varying environmental

conditions, including those of pH, temperature, oxidation-reduction potential, soil type,

and hydrology (Moore and Ramamoorthy, 1984; Lodenius et al 1987; Del Debbio, 1991;

Barrow and Cox, 1992b). Anthropogenic activities may increase mercury inputs to the

environment or they may elicit the transport or transformation of ambient mercury.

Presently, anthropogenic mercury inputs comprise approximately one-half of the

mercury entering the world ecosystem (Fitzgerald and Clarkson, 1991). There are

indications that atmospheric mercury concentrations are steadily increasmg (Slemr and








9

Langer, 1992) and increased rates of mercury accumulation to aquatic systems have been

demonstrated in the sediment record (Meger, 1986; Norton et al., 1990; De Lacerda et

al., 1991; Swain et al., 1992). Elevated mercury concentrations in fish pose a human

health hazard and mercury biomagnification through the food web provides evidence of

the ecological stresses imposed by this trace metal (Cardeilhac et al., 1981; Hand and

Friedemann, 1990; Roelke et al., 1991; Heaton-Jones, 1992; Jurczyk, 1993). Analytical

technology has been challenged by the unique problems (i.e. low concentration, low vapor

pressure, analytical contamination, speciation, toxicity, and biomagnification) associated

with environmental mercury research (Fitzgerald, 1986; Schroeder, 1989; Douglas, 1991).


Human-Related Activities


Mercury has been used widely: 1) for the production of electrical devices, 2) as

a catalyst for the chlor-alkali industry, 3) as a fungicide/algicide in paint products,

paper/pulp manufacture, and agriculture, and 4) as a component in the manufacture of

instruments (i.e. thermometers), dental preparations, and pharmaceuticals (Mitra, 1986;

Nriagu, 1990; KBN Engineering and Applied Sciences, Inc., 1992). Anthropogenic

releases of mercury to the environment are related to activities including the burning of

fossil fuels (Crockett and Kinnison, 1977; Sengar et al., 1989; Lodenius, 1990), the

incineration of municipal solid and medical waste (Collins and Cole, 1990; Volland, 1991;

KBN Engineering and Applied Sciences, Inc., 1992), the production of electricity

(Lindberg, 1980), wastewater discharge (Morel et al., 1975), agricultural practices

(Simons, 1991; Patrick et al., 1992), mining (Pfeiffer et al., 1991), and chlor-alkali and








10

cement manufacture (Fukuzaki et al., 1986; Mitra, 1986). Further, land development and

hydrologic manipulation facilitates the release of naturally derived mercury from deep

("old") sediment (Horvath et al., 1972; Simola and Lodenius, 1982). Collins and Cole

(1990) outlined a mass balance of mercury discharges to the environment (Table 2.1).





Table 2.1. Human-related discharges of mercury to the
environment (Kg yr').

Source 1973 1988

Industry
Chemical manufacture 307,709 18,043
Petroleum refining 36,943 227
Smelting 65,743 0
Electronics manufacture 185,394 1,000

Utilities
Coal burning 40,625 73,483
Natural gas 27,393 N/A

Incinerators 16,829 40,234









There is a rich, centuries-old, history of the contribution that mercury has played

in society (Fitzgerald, 1986; Mitra, 1986). Environmental mercury contamination was

initially identified in response to human tragedy, such as mass poisoning and death (i.e.

Minimata disease)(D'Itri, 1991), born of the careless use of mercury and it's haphazard









11

disposal (Horvath et al., 1972: Hamdy and Post, 1985; Collins and Cole, 1990)

Subsequent observations of the deleterious effect of mercury (Hakanson et al., 1990a,

1990b; Scheuhammer, 1991a, 1991b) on the world ecosystem clearly identified the need

to establish stringent guidelines for the use of mercury-containing compounds in industry

and agriculture (Revis et al., 1990; Ingersoll, 1991). However, population growth and

development pressures have created new avenues by which society may contribute to the

mercury budget of the world ecosystem through the burning of fossil fuels, medical and

municipal solid waste incineration, and electricity production (Horvath et al., 1972;

Albrinck and Mitchell, 1979; Collins and Cole, 1990).


Mercury Issues in Florida


Widespread mercury contamination was identified in Florida after the discovery

of elevated mercury concentrations in fish throughout the state (Hand and Friedemann,

1990). The death of an endangered Florida panther was attributed to mercury toxicosis

(Roelke et al., 1991), and mercury accumulation was cited as a potential cause for

dramatic declines of wading bird populations (Jurczyk, 1993). A mercury emissions

survey identified municipal solid waste (MSW) and medical waste incineration, the

electric utility industry, and paint application as the primary anthropogenic sources of

atmospheric mercury emissions in Florida (KBN Engineering and Applied Sciences, Inc.,

1992), and agricultural practices have been identified as potential release mechanisms for

naturally and anthropogenically derived mercury reserves in rich organic soils in the state

(Simons, 1991).











Available Technology for Mercury Research


Mercury concentrations in water and air are much less than those found m soil,

sediment, plant tissue and animal tissue (Schroeder, 1989). To understand and evaluate

the environmental impacts of mercury contamination, and the cycling of mercury in the

ecosystem, analysts have been faced with a stiff technical challenge (Douglas, 1991).

Ambient mercury concentrations in water and air often fall near, or below, the limits of

detection provided by many traditional analytical techniques (Bloom, 1989; LeBihan and

Cabon, 1990). Contamination during sampling, storage, and analysis of such samples may

exceed actual ambient mercury concentrations (Fitzgerald and Watras, 1989). Mercury

transport, bioaccumulation, and toxicity in the environment often depends on

environmental conditions and mercury speciation (Cope et al., 1990; Farrell et al., 1990;

Lodenius, 1990; Verta, 1990; Johnston et al., 1991; Nilsson and Hakanson, 1992). As a

consequence, the constraints imposed by environmental mercury studies challenge

researchers to optimize the available analytical technology to provide a suitable basis with

which to characterize mercury abundance and transformation in the environment.

The sample matrix and mercury content must be considered when selecting an

analytical procedure to determme mercury in environmental samples. The selected

technique must be sufficiently sensitive to quantify the anticipated mercury concentration

and must demonstrate robustness when challenged by matrix interference inherent to

particular sample types (air, water, soil, biota). Separation and speciation issues

associated with a given sample matrix must also be addressed (Schroeder, 1989).

Cold vapor atomic absorption spectrophotometry (CVAAS) has been the standard

analytical method for mercury determinations in environmental and biological samples








13

(Winter et al., 1977; Perry et al., 1978). Numerous modifications have been developed

to decrease sampling time, to increase analytical sensitivity (Freimann and Schmidt, 1982;

Mateo et al., 1990; Munaf et al, 1990a; Welz et al., 1992), and to facilitate mercury

speciation (Schroeder, 1989; Munaf et al., 1990b; Rapsomanikis and Craig, 1991; Craig

et al, 1992;).

Various preconcentration steps, such as mercury-gold amalgamation (Freimann and

Schmidt, 1982), continuous flow, and on-line pretreatment have been used to improve

sensitivity and efficiency (Mateo et al., 1990; Munaf et al., 1990a; Welz et al., 1992).

On-line (Munaf et al., 1990b; Rapsomanikis and Craig, 1991; Craig et al., 1992) and off-

line (Schroeder, 1989) separation techniques have been employed to facilitate mercury

speciation.

A variety of alternative techniques have been used to improve mercury detection

and speciation. Voltammetric techniques have been enhanced by preconcentration

strategies (Daih and Huang, 1992) and electrode modification (Navratilova and Kula,

1992). Electrothermal atomization atomic absorption spectrophotometry has been used

after solvent extraction preconcentration (LeBihan and Cabon, 1990). Mercury speciation,

using chromatographic separations by high performance liquid chromatography

(HPLC)(Krull et al., 1986) and capillary gas chromatography (Kato et al., 1992), followed

by atomic emission detection has been described. Gas chromatographic (GC) techniques

for methylmercury determination traditionally used electron capture detection (Horvat et

al., 1988) because of the sensitivity of the detector. Recently studied GC techniques

employ sample preconcentration (Lansens et al., 1990; Bulska et al., 1991) or headspace








14

injection (Lansens et al., 1989) coupled with microwave-induced plasma (MIP) detection

(Lansens et al., 1989; Lansens et al, 1990; Bulska et al., 1991) and inductively coupled

plasma-mass spectrometry (ICP-MS)(Shum et al., 1992).

Mercury analyses in water and air historically have been flawed by contamination

that often exceeds ambient mercury concentrations (Fitzgerald and Watras, 1989).

Technical advances have incorporated new strategies for sampling (i.e. "clean sampling

technique" )(Douglas, 1991), and detection limits have been lowered by implementation

of clean laboratory practices and improved analytical techniques (i.e. atomic fluorescence

spectrophotometry)(Bloom, 1989). Further, improved technology has enabled researchers

to quantify individual mercury species (Fitzgerald, 1986).

Atomic fluorescence spectrophotometry is a most promising and versatile

technique for mercury detection in environmental matrices, and is rapidly becoming

accepted as the standard technique for low-level mercury determinations (Bloom, 1989;

Tanaka et al., 1992). Fluorescence technology is free from spectral interference that

plague absorption technology (Churchwell et al., 1987). Improved mercury detection

limits, furnished by cold vapor atomic fluorescence spectrophotometry (CVAFS), approach

0.6 pg Hg (0.003 ng L-' for a 200 mL sample)Bloom, 1989). Basic fluorimetric

spectrophotometry (Mariscal et al., 1992) has been used to improve the sensitivity of total

mercury determinations, while atomic fluorescence, following preconcentration and

chromatographic separation (Bloom, 1989; Tanaka et al., 1992) permits mercury

speciation at very low analyte concentrations. Extensive speciation schemes that employ

"clean field and laboratory" procedures (Douglas, 1991), and improved separation and








15

analytical techniques (Wilken, 1992) broaden our ability to quantify mercury in

environmental matrices and to identify ecological transformations of mercury.

Improvements in atmospheric mercury determinations have incorporated

concentration steps, such as gold trap amalgamation (Barghigiani et al., 1991), or selective

absorption tubes, to permit mercury speciation (Braman and Johnson, 1974; Schroeder and

Jackson, 1987). Neutron activation analysis after preconcentration (Albrinck and Mitchell,

1979) and LIDAR techniques have been described (Ferrara et al., 1992).

The detection limits provided by traditional technology, such as cold vapor atomic

absorption spectrophotometry (CVAAS), have typically been sufficient for the

determination of mercury in soil, sediment and biological samples (Sullivan and Delfino,

1982; Colina de Vargas and Romero, 1992). Systematic mercury contamination during

sampling and analysis does not usually influence the quantification of mercury in these

matrices. Microwave digestion (Navarro-Alarcon et al, 1991) and gold amalgamation

preconcentration (Mudroch and Kokotich, 1987), respectively, speed sample preparation

and improve sensitivity of CVAAS technology. Separation techniques have been

employed to evaluate certain mercury species in environmental and biological samples.

For example, methylmercury can be determined, after solvent extraction and subsequent

identification/quantification, using gas chromatography with electron capture detection

(GC-ECD)(Alvarez and Hight, 1984; Hight, 1987; Horvat et al., 1990; Bulska et al.,

1991). While modifications of the GC-ECD method have employed improved extraction

procedures and analytical configurations (Lansens and Baeyens, 1990), organomercurials

have also been characterized with methods using HPLC (Hempel et al., 1992; Stoeppler

et al, 1992) and ICP-MS (Beauchemin et al., 1988) technology.








16

Another approach to mercury speciation is to establish operational definitions that

categorize mercury species based on a common response to a series of physicochemical

conditions (Schroeder, 1989). According to these speciation schemes, compound groups

are isolated by a variety of sequential selective extraction procedures (Magos, 1971; Revis

et al., 1990; Rapsomanikis and Andreae, 1991; Sakamoto et al., 1992).

Much research must follow guidelines that are outlined by state or federal agencies

(Winter et al., 1977). As a consequence of regulated adherence to "standard methods,"

researchers are often limited by traditional analytical technology until the regulatory

agency accepts modified and contemporary technology. The expanding technological

advances for trace metal analyses, and the complexity associated with environmental

analytical chemistry (i.e. variable analyte concentration, speciation, and matrix

interference) necessitate that the researcher: 1) optimize a protocol for "self-evaluation"

in the laboratory, and 2) implement interlaboratory calibration studies that characterize the

utility of traditional and contemporary techniques when analyzing environmental matrices.

Comparative studies of parallel methods (Churchwell et al., 1987; Horvat et al., 1988;

Friese et al., 1990) and interlaboratory calibration studies (Thibaud and Cossa, 1989;

Cossa and Courau, 1990) allow researchers to: 1) compare methods and optimize routine

laboratory practices, 2) identify superior analytical activities (sampling, storage,

preparation, and analysis) and, 3) evaluate the quality of data provided by a particular

method or laboratory.

Cold vapor atomic absorption spectrophotometry (CVAAS) is suitable for total

mercury determinations in soil, sediment and biological tissue if measurements are








17

verified by the appropriate quality assurance/quality control measures (i.e. instrument

calibration against a standard reference material, and verification of instrument stability).

However, sample preconcentration or analytical modifications are essential for mercury

determinations of air and water samples. Low level mercury determinations in water and

air samples, using CVAAS technology, should be considered suspect until these data can

be compared with external determinations using alternative technology.


Global Mercury Cycle


The global cycle of mercury is mechanistically determined by its high vapor

pressure (2.4 x 10-3 mm Hg at 200C)(Stewart and Bettany, 1982; Clarkson et al., 1984;

Schroeder et al., 1989). This unique physico-chemical attribute causes the global mercury

cycle to be distinctly different from that of other trace metals (Moore and Ramamoorthy,

1984). The global cycle of mercury, involving the solid, aqueous, and vapor phases, and

influenced by the stability of volatile mercury species, permits widespread and long-term

dispersion of this element. The global cycle of mercury is outlined in Figure 2.1.

Mercury is released to the atmosphere from natural land and ocean degassing,

volcanic activity, and human-related activities. Particulate-phase mercury is deposited

readily from the atmosphere, while vapor-phase mercury enters the global atmospheric

cycle and is dispersed for long distances. Photo-oxidative processes and particulate-

scavenging mechanisms eventually convert the vapor-phase mercury into a particulate

form that is deposited by dry deposition or is scavenged during precipitation events.




















Biosphere



A B


H

G


F



E


Assimilation
Decay
Decay
Assimilation
Metamorphism
Dissolution
Assimilation


Decay
Weathering
Mineralization
Volcanism
Evaporation
Condensation
Volcanism


Figure 2.1. Block Diagram of the Global Mercury Cycle


Water


K
--








19

Mercury is transported from the land to aquatic environments by terrestrial

leaching or by discharges associated with human-related activities. Sediment serves as

the primary sink for mercury as a result of the strong affinity of mercury for organic and

sulfidic substrates, although a fraction (<1%) of sediment mercury may be remobilized

as labile mono- or dimethylmercury. Monomethylmercury biomagnifies in the food chain

and volatile dimethylmercury is released to the atmosphere. Natural terrestrial and

oceanic releases of mercury to the atmosphere (30-100 x 108 g Hg yr1 and 20-100 x 108

g Hg yr', respectively) are roughly equivalent to anthropogenic atmospheric releases (20-

100 x 108 g Hg yr-') (Fitzgerald, 1986; Kim and Fitzgerald, 1986).

Atmospheric mercury deposition to the terrestrial environment is estimated to be

(40-100) x 108 g Hg yri, while mercury deposition on the world ocean is estimated to be

(20-275) x 108 g Hg yr-' (Fitzgerald, 1986). The broad estimates for the global mercury

cycle arise from the sparsity of reliable data for certain compartments of the environment

(Fitzgerald, 1986; Fitzgerald and Clarkson, 1991).


Global and Regional Interactions


In the atmosphere, particulate-phase mercury may be transported in a manner

similar to other metals (Nater and Grigal, 1992). For example, particulate emissions from

point sources such as volcanoes, fires, and industry, may establish regional mercury

gradients in the surrounding environment (Crockett and Kinnison, 1977; Lindberg, 1980;

Fukuzaki et al., 1986; Sengar et al., 1989; Barghigiani and Ristori, 1991; Pfeiffer et al.,

1991; Ferrara et al., 1992). The transport of particulate-phase atmospheric mercury








20

depends on wind direction (Brosset, 1987). However, more than 95% of the total

atmospheric mercury inventory is in the gaseous elemental form, with an atmospheric

residence time of 0.7 to 2.0 years (Slemr and Langer, 1992).

Since >95% of the total atmospheric mercury inventory is in the gaseous elemental

form, mercury accumulation in regional terrestrial and aquatic systems may be dictated

by global changes in the mercury cycle (Swain et al., 1992). Conversely, local activities

may contribute readily to the global cycle.

Mercury deposition in Swedish soils has been linked to mercury emissions from

the United Kingdom, Germany, and Poland (Hakanson et al., 1990b) and 10-15% of the

mercury in fish from Swedish lakes has been attributed to mercury emissions from foreign

sources (Hakanson et al., 1990a). Recent studies have attributed increases in sediment

mercury accumulation to increased global atmospheric mercury emissions (Meger, 1986;

Steinnes and Andersson, 1991; Swain et al, 1992) corresponding to an estimated 2%

annual increase in the atmospheric mercury budget (Slemr and Langer, 1992). Natural

inputs of mercury to the global cycle include volcanism (Barghigiani and Ristori, 1991),

tectonic activity (Varekamp and Waibel, 1987), and ocean and land degassing (Xiao et

l, 1991).


Mercury in the Atmosphere


The total gaseous mercury (TGM) concentration (>99% Hg) comprises greater

than 95% of the total atmospheric mercury component (Bloom and Watras, 1989). Total

gaseous mercury concentrations range between 1 and 7 ng m" for samples taken from the








21

Pacific Ocean, the Mediterranean Sea, Italy, and rural Wisconsin (Fitzgerald et al., 1984;

Ferrara et al., 1986; Fitzgerald et al., 1991)(Table 2.2). Given the sparsity of data, there

is no evidence for continental sources of TGM. Elemental mercury in the atmosphere has

a relatively long residence time (0.7 to 2.0 years)(Munthe and McElroy, 1992; Slemr and

Langer, 1992) and ambient concentrations are not significantly influenced by rain episodes

(Ferrara et al., 1986). Elemental mercury is eventually oxidized by a variety of chemical

oxidations (i.e. ozonation)(Iverfeldt and Linqvist, 1986) and photo-oxidative processes

(Munthe and McElroy, 1992) and the resulting ionic mercury species (Hg" and CH3Hg+)

are readily scavenged by rainfall (Iverfeldt and Linqvist, 1982). During rain episodes,

there is a washout event that delivers water-soluble mercury to the earth's surface (Ferrara

et al., 1986). Wet deposition is not a significant source of monomethylmercury to the

equatorial Pacific Ocean (Mason et al., 1992) and a Wisconsin seepage lake (Fitzgerald

et al., 1991), however, Bloom and Watras (1989) suggest that monomethylmercury

concentrations of 0.15 ng L'1 ([Hg]1,,= 2-5 ng L') in precipitation in the northwestern

United States can account for most of the fish mercury budget in Washington state lakes.



Mercury in Water


Mercury can occur in natural waters in the elemental (Hg), mercurous (Hg*'), or

mercuric (Hg+2) forms, depending on ambient pH, oxidation-reduction potential, and ionic

composition. The thermodynamic stability domains (EH-pH diagram) for predominant

compounds, under varying pH and redox conditions, are described in Figure 2.2.

Mercuric hydroxy- and chloro- complexes are favored under conditions of high ambient











Table 2.2. Atmospheric mercury concentrations and stack emission concentrations.


Industry/Source [Hg], ng mi3 References


Point Sources Emissions
-Mercury smelter
-Chlor-alkali plant
-Coal-fired power plant
-Coal-fired power plant
-Coal-fired power plant
-Non-ferrous smelter
-Sewage sludge incinerator
-Cement factory
-Volcano (Mt. Etna, Sicily)
- Cinnabar deposit, (Mt. Amiata, Italy)

Ambient Atmospheric Concentrations
- Italy and Mediterranean Sea
Capraia Island (sea level)
San Pelligrinetto, Italy (1000 m)
Livomo, Italy (urban)
R. Solvay, Italy (chlor-alkali)
Mt. Amiata (cinnabar deposits)
Mt. Amiata (10-20 m above ground)
Equatorial Pacific Ocean
North Central Pacific Ocean
Little Rock Lake, WI


Albrinck and Mitchell, 1979
Albrinck and Mitchell,1979
Germani and Zoller, 1988
Albrinck and Mitchell. 1979
Lindberg, 1980
Albrinck and Mitchell, 1979
Albrinck and Mitchell, 1979
Fukuzaki et al., 1986
Barghigiani and Ristori,1991
Ferrara et al., 1992


6.8
5.7
10.1
22.5
16.4
2.5
1.3
1.8
1.6


Ferrara et al., 1986
Ferrara et al., 1986
Ferrara et al., 1986
Ferrara et al., 1986
Ferrara et al., 1986
Ferrara et al., 1986
Ferrara et al., 1992
Fitzgerald et al., 1984
Fitzgerald et al., 1991
Fitzgerald et al., 1991


Precipitation [Hg], pM References

Italy and Mediterranean Sea Ferrara et al., 1986
- Capraia Island (sea level) 96 Ferrara et al., 1986
- San Pelligrinetto, Italy (1000 m) 50 Ferrara et al., 1986
- Livomo, Italy (urban) 133 Ferrara et al., 1986
- R. Solvay, Italy (chlor-alkali) 131 Ferrara et al., 1986
- Mt. Amiata (cinnabar deposits) 100 Ferrara et al., 1986
Northeast Pacific Ocean 45 Fitzgerald et al., 1991
Wisconsin, USA 52 Fitzgerald et al., 1991
Washington, USA 17 Bloom and Watras, 1989












1.2

1.0

0.8

0.6

0.4

0.2

0.0

0.2
0.4

0.6


0 2 4 6 8 10 12 14


Figure 2.2: Thermodynamic stability diagram for mercury
(redrawn from Krabbenhoft and Babiarz, 1992)


0?
0

-c

w








24

pH and chloride concentrations. Mercury is readily completed by high molecular weight

dissolved organic materials humicc and fulvic acids), typically associated with organic

sulfhydryl moieties (Andren and Harriss, 1973; Mantoura et al., 1978). Mercuric sulfides

are favored in reducing environments (Lodenius et al., 1987).

Mercury is delivered to aquatic systems from direct precipitation and terrestrial

runoff (Xiankun et al., 1990). Typical mercury concentrations in fresh, estuarine, and

saline waters are presented (Table 2.3). In aquatic systems, mercury is readily adsorbed

to the surface of living and nonliving particulate material (Wilkinson et al., 1989) due to

the strong adsorption capacity of organic particulates for mercury (Bilinski et al., 1992).

Partition coefficients for mercury between suspended solids and water have been

calculated to be (1.34 1.88) x 10' (Moore and Ramamoorthy, 1984). These adsorption

and complexation processes increase the rate of mercury removal to the sediment via

particulate scavenging and sedimentation (Mantoura et al., 1978; Wallace et al., 1982;

Moore and Ramamoorthy, 1984). During estuarine mixing, increased salinity induces the

precipitation of mercury-humic complexes, and in saline environments, mercuric chloride

complexes may become a dominant mercury component (Morel et al., 1975; Calmano et

al., 1992).

The low concentrations of mercury in natural waters necessitate the use of clean

sampling techniques and contemporary analytical technology (Bloom, 1989; Douglas,

1991). Many traditional sampling and analytical techniques for mercury in natural waters

are confounded by errors due to contamination and/or analytical insensitivity (Fitzgerald

and Watras, 1989), hence, caution must prevail when evaluating historic data.











Table 2.3. Mercury concentrations determined for various natural water bodies.


Location Type [Hg], pM Hg species References


Swedish lakes


Little Rock Lake, WI
Gironde Estuary, France
St. Lawrence Estuary,
Canada
Pacific Ocean
Alboran Sea, Spain
Strait of Gibraltar, Spain
East North Atlantic
English Channel, England
Bay of Biscay, France

Nova Scotia

Adriatic Sea, Italy

North West Atlantic
North Central Pacific
North Atlantic


North Pacific

Equatorial Pacific



Baltic Sea, Germany


FW
FW
FW
FW
EST
EST
SW
SW
SW
SW
SW
SW
SW
SW
SW
SW
SW
SW
SW
SW
SW
SW
SW
SW
SW
SW
SW
SW
SW
SW


0.8 2.0
0.6- 1.3
0.5 0.6
0.7- 2.9
21.8 103.2
9.0- 15.0
2.4
4.7 9.7
0.2 0.7
0.2 0.6
0.4- 10.0
1.0 20.4
2.8 4.3
1.4 2.8
2.2 (near-shore)
2.3 (off-shore)
10.1 33.7
0.4 76.5
3.3 4.7
1.7- 2.5
4 (surface)
10 (thermocline)
<4 (deep)
1.2 2.6
1.0 1.8
0.3 5.0
0.1 1.0
0.0 0.6
0.0 0.7
2.5


References:






Abbreviations:


'Lee and Hultberg, 1990; 2Fitzgerald and Watras, 1989;
'Cossa and Noel, 1987; 4Cossa et al., 1988; 'Cossa and Martin, 1991;
6Cossa et al., 1988; 7Cossa and Fileman, 1991; 8Dalziel, 1992;
'Ferrara and Maserti, 1992; 'oGill and Fitzgerald, 1987;
"Gill and Fitzgerald, 1988; '2Mason and Fitzgerald, 1990;
'3Schmidt, 1992
SW (saline water), EST (estuarine water), FW (fresh water),
DGM (dissolved gaseous mercury), MMHg (monomethylmercury),
DMHg (dimethylmercury), total and reactive (total and reactive mercury)


methyl-Hg
methyl-Hg
methyl-Hg
reactive
total
dissolved
dissolved
total
reactive
reactive
total
total
total
reactive
reactive
reactive
dissolved
particulate
total
total
total
total
total
total
total
reactive
DGM
MMHg
DMHg
total










Mercury in Sediment


Sediments are a primary sink for mercury in the environment (Tolonen et al.,

1988). Mercury concentrations found in contaminated and noncontaminated soil and

sediment are presented (Table 2.4). Mercury forms strong associations with organic

maternal in soil and sediment under aerobic conditions. In addition, under anaerobic

conditions, insoluble mercuric sulfides may form (Lindberg and Harriss, 1974). Lindberg

and Harriss (1974) found that mercury in sediment porewater was associated with low

molecular weight (MW<500) dissolved organic matter in Florida Everglades sediment, and

with high molecular weight (MW>100,000) dissolved organic matter in sediment from

Mobile Bay, Alabama.

Senaratne and Dissanayake (1989) hypothesized a mechanism by which dissolved

mercury, initially scavenged from the water column by organic particulate material, is

precipitated as mercuric sulfide after reducing conditions are established in response to

the decomposition of sedimented organic material. Their estuarne studies further

suggested that mercury was readily adsorbed to the surface of detrital grains coated with

iron-manganese oxides. Controlled Experimental Ecosystem studies (Wallace et al., 1982)

demonstrated the rapid removal of mercury from the water column, where more than 90%

of spiked mercury was associated with particulate, colloidal, and high molecular weight

organic materials in the sediment. Winfrey and Rudd (1990) added 203Hg to an organic

sediment, demonstrating that more than 99% of the radio-labelled mercury was retained

by the substrate. Likewise, Krabbenhoft and Babiarz (1992) found that 92% to 96% of

deposited mercury was retained by soils. The retention of mercury by organic soils is




















Table 2.4. Mercury concentrations in variously impacted soil and sediment.


Location [Hg], mg Kg-' References

Little Rock Lake, WI sediment 0.10 1
Canadian peat 0.06 2
Okefenokee, GA peat 0.40 2
Lake Istokpoga, FL peat 2.45 2
Swedish soils 0.12 0.22 3
Sri Lanka tidal flat 4.40 4
Sri Lanka peat sediment 15.5 4
W. Everglades mangroves 0.22 1.86 5
Adriatic Sea 0.02 8.63 6
Municipal Wastewater Sludge 1.24 7
Sewage Treatment Plant (STP) study*
Returned Activated Sludge (RAS) 12.3 30.0 8
Non-Hg contaminated RAS 0.03 8
Sediment (upstream of STP) 0.09 8
Sediment (downstream of STP) 0.98 8
Sewage-amended soil 14.6 8
Sediment (chlor-alkali receiving lagoon) 420. 8

'Fitzgerald and Watras, 1989; 2Roddy and Tomlinson, 1989; 3Steinnes and Andersson, 1991;
4Senaratne and Dissanayake, 1989; 'Lindberg and Harriss, 1974; 6Ferrara and Maserti, 1992;
'Cappon, 1984; 'Olson et al_, 1991








28

much stronger that for mineral soils due to the association of mercury with sulfhydryl

moieties in the organic materials (Barrow and Cox, 1992a, 1992b).

Barrow and Cox (1992a, 1992b) characterized the influence of ambient salinity on

mercury sorption. At low chloride concentrations, mercury sorption to organic material

was unchanged between pH 4 and 6, and decreased at pH values greater than 6 (Barrow

and Cox, 1992a). The maximum sorption of mercury to the mineral, geothite, occurred

at a pH less than 4 (Barrow and Cox, 1992b). At high chloride concentrations, mercury

sorption to organic material increased between pH 4 and 6, and decreased at pH greater

than 6 (Barrow and Cox, 1992a).

Complexation of mercury in humic-rich or saline waters may facilitate the

desorption of mercury from sediment (Lindberg and Harriss, 1974). Studies in the Krka

estuary of Russia suggested that mercury, associated with dissolved organic material in

the freshwater environment, was readily precipitated during estuarine mixing, and that

mercury adsorption to mineral surfaces enhanced sedimentation at the freshwater/saltwater

interface (Bilinski et al., 1992). Mercury removal during estuarine mixing likely

decreased mercury bioavailabilty in these regions (Calmano et al., 1992).

Mercury from deep sediment may be released to overlying water in particulate

form as a result of deep sediment cracking that can occur during repeated drying and

flooding events (Lodenius et al., 1987). Desorption of mercury from sediment can occur

under acidic conditions; however, typical sediment pH values are not sufficiently low to

elicit this response (Barrow and Cox, 1992a, 1992b). Strong bonds, between mercury and

sulfhydryl moieties of particulate organic matter, render mercury unavailable for uptake








29

(Langston, 1982, Duddridge and Wainwright, 1991). Further, less than 1% of the

inventory of sediment mercury is present as methylmercury, thus largely unavailable for

assimilation (Hennig et al., 1989; Revis et al., 1990; Winfrey and Rudd, 1990).


Mercury in Biota


Various mercury levels have been identified in plant and animal tissues (Table

2.5). Mercury was shown to accumulate in the root tissue of Spartina altemiflora but was

not readily transported to the rhizomes and above-ground tissues (Breteler et al., 1981).

Further, mercury accumulation rates did not increase for plants grown in soils that were

amended with sewage sludge fertilizers, but accumulation was inversely related to

sediment organic matter content (Breteler et al., 1981). Elevated temperature and light

intensities increased mercury uptake in the aquatic macrophytes, Elodea densa (Maury-

Brachet et al., 1990) and Ludwigia natans (Ribeyre, 1991). Fortmann et al. (1978)

suggested that plant mercury uptake can make deep sediment mercury available for

accumulation in the food chain or facilitate its return to the global cycle in the aqueous

or gas phases, or as detritus. Mercury abundance has been measured in plankton (Watras,

1993), invertebrates, fish (Barber and Whaling, 1984; Lange et al., 1993), birds (Burger

and Gochfeld, 1991; Thompson et al., 1992), raccoons (Roelke et al., 1991), and top

carnivores (i.e. American alligator and Florida panther)(Roelke et al., 1991; Heaton-Jones,

1992).

Historically, most studies evaluated mercury accumulation in animals because of

human health concerns. As a result, studies of mercury in consumables, such as sportfish,

dominate the literature (Schmitt and Brumbaugh, 1990). Growing ecological concerns













Table 2.5. Mercury concentrations in plant and animal tissues.

Sample [Hg], mg/Kg References


Marine Fish/Mammals
-Sardines
-Dolphins (Equat. Pacific Ocean)
-Pacific Blue Marlin
-Snapper
-Fish/Shellfish (Minamata, Japan)
Freshwater Fish
-Brown Trout
-Northern Pike
-Perch (Wisconsin lakes)
-Bass (Florida lakes)
-United States survey

-Walleye (Manitoba reservoir)"
-Walleye (Manitoba reservoir)b
Alligator (Florida Everglades)
-Farm-raised
-Native
Florida Panther (Florida Everglades)
Plankton
-Adriatic Sea
-Unknown
-Spain
Birds
-Starlings (United States survey)
Plants
-Lichens (Yugoslavia)
-Lichens (Finland)
-Fungi (Finland)
-Mosses (Finland)
-Ferns (Finland)
-Pines (Finland)
-Angiosperms (Finland)
-Angiosperms (USA)'


0.02
1. -
14.0
0.01 -
9.

0.08
27.8
0.06 -
0.03 -
0.01 -

0.2 -
0.5 -


5


1.66
24



0.19
1.38
0.37


0.3
1.0


0.10
1.50
100 (liver)


0.02 -
0.02 -
0.50 -


Beckert, 1978
Andre et al., 1990
Beckert, 1978
Chvojka et al., 1990
Beckert, 1978

Beckert, 1978
Beckert, 1978
Cope et al., 1990
Lange et al., 1993
Schmitt and Brumbaugh,
1990
Bodaly et al., 1984
Bodaly et al., 1984

Heaton-Jones, 1992
Heaton-Jones, 1992
Roelke et al., 1991

Ferrara and Maserti, 1992
Mitra, 1986
Mitra, 1986

White et al., 1977

Lupsina et al., 1992
Nuorteva et al., 1986
Nuorteva et al. 1986
Nuorteva et al., 1986
Nuorteva et al., 1986
Nuorteva et al., 1986
Nuorteva et al., 1986
Mitra, 1986


0.14
0.04
16.80


0.01 0.20


0.40-
0.06 -
0.05 -
0.04-
0.01 -
0.03 -
0.01 -
0.5 -


188.2
0.57
1.40
0.67
0.06
0.15
0.22
3.5


"before flooding of reservoir,
after flooding of reservoir,
'trees growing above cinnabar deposit








31

have resulted in the expansion of the scope of organismal mercury studies. Some studies

are geared toward understanding the role of food chain dynamics on biomagnification

(Watras, 1993). Florida panthers, for example, that fed primarily on raccoons exhibited

higher tissue mercury concentrations than those that fed on deer (Roelke et al., 1991).

Watras (1993) studied mercury in zooplankton from Little Rock Lake, Wisconsin, and

suggested that increased methylmercury production, resulting from lake acidification,

resulted in increased mercury accumulation, and that bioconcentration factors (BCF) for

methylmercury species were related to the trophic level of the test organism.


Environmental Factors and Bioaccumulation


Mercury Transformations in Aquatic Systems


Methylation of mercury in the water and sediment of aquatic environments has

been shown to enhance the bioavailability of mercury to biota (Cope et al., 1990; Watras,

1993). Numerous studies have generated a wealth of information regarding methylation

and demethylation of mercury in aquatic systems. However, the mechanistic complexity

of the aquatic mercury cycle confounds many attempts to characterize exclusive cause-

effect relationships among habitats (Miskimmin et al., 1992).

The rate of mercury methylation is optimized under acidic and freshwater

conditions at an ambient temperature around 350C (Bryan and Langston, 1992). Rates

of methylation decrease with increased pH and salinity, and are favored under moderately

anoxic conditions. Since methylmercury comprises less than 0.2% of the total mercury

concentration, it is unlikely that methylation plays a significant role in the post-








32

depositional migration of mercury, although methylmercury has been shown to play a

significant role in mercury bioavailability (Bryan and Langston, 1992)

Increased mercury release from anoxic sediment during lake stratification, with

subsequent precipitation by free sulfides, follows a pattern that is typical of the redox-

active metals (Bloom and Effler, 1990). However, the microbial (Berman et al, 1990;

Choi and Bartha, 1993) and abiotic (Ebinghaus and Wilken, 1993) methylation of mercury

in water and sediment adds another level of complexity to the mercury cycle.

Monomethylmercuric ion is readily assimilated into living tissue due to the propensity of

the mercuric ion to bind with sulfhydryl moieties of organic compounds (Hintelmann et

al., 1993). However, competitive mechanisms of assimilation, monomethylmercuric

sulfide precipitation, and sulfide-mediated disproportionation of monomethylmercuric ion

to volatile dimethylmercury, participate in a dynamic disequilibrium that influences the

compartmentalization of mercury between sediment, water, and biota (Bloom and Effler,

1990; Ferrara and Maserti, 1992). This disequilibrium is further driven by a variety of

environmental conditions including pH, alkalinity, organic content, and oxidation-

reduction potential (Steffan et al., 1988; Winfrey and Rudd, 1990; Miskimmin et al

1992). For example, alkaline/anoxic conditions favor sediment dimethylmercury

production (Quevauviller et al., 1992).

Abundant quantities of dissolved organic material diminish methylation due to

mercuric ion chelation (Miskimmin et al., 1992). Methylmercury production increases

with decreasing pH from 7 to 5 at the sediment-water interface (Winfrey and Rudd, 1990;

Miskimmm et al_, 1992), but decreases with decreasing pH in anoxic/subsurface sediment








33

(Steffan et al., 1988; Winfrey and Rudd, 1990). Volatilization of elemental mercury

decreases with decreasing pH (Winfrey and Rudd, 1990) and does not exceed 2% of the

ambient methylation activity.

Descriptive examples of the complexity of the mercury cycle in aquatic systems

include some antagonistic mechanisms imposed by physical and microbial conditions.

Miskimmin (1991) demonstrated that the solubility of monomethylmercury was directly

related to the dissolved organic carbon (DOC) content in natural waters. Although

sediment mercury methylation was not related to the DOC of overlying water, the ambient

DOC facilitated the release of available monomethylmercury to overlying waters.

Interestingly, monomethylmercury(II)-DOC complexes were shown to diminish mercury

bioavailability.

Sulfate reducing bacteria have been identified as key participants in the

methylation of mercury (Berman et al, 1990; Winfrey and Rudd, 1990; Oremland et al.

1991; Choi and Bartha, 1993). While providing the mechanism to enhance

bioaccumulation via methylation, sulfate-reducing bacteria produce sulfide, as a by-

product of respiration. Ambient sulfides, produced by microbial sulfate reduction, may

effectively immobilize labile monomethylmercuric ion as monomethylmercuric sulfide

(Bloom and Effler, 1990). Interestingly, although more than 95% of methylation results

from cobalamin-mediated methylation by sulfate-reducers (Berman et al., 1990),

methylation rates were shown to increase under sulfate-limiting conditions (Berman et al.,

1990; Choi and Bartha, 1993). Under controlled condition, sulfate-reducing bacteria

methylated less than 1% of available mercury under sulfate-reducing conditions and








34

methylated approximately 40% of available mercury under sulfate-limiting, fermentative

conditions.

Finally, demethylation processes provide another pathway for mercury speciation.

Monomethylmercury may be demethylated via an organomercurial lyase pathway in which

the covalent carbon-mercury bond is cleaved enzymatically, and the resultant mercuric ion

is reduced to elemental mercury with an enzyme-mediated mercuric reductase pathway

(Nakamura et al., 1990). Alternatively, methylmercury may undergo an oxidative

demethylation, in which monomethylmercury is used as an analog of a single-carbon

substrate for metabolism, with the concomitant production of carbon dioxide (Oremland

et al., 1991).

The competitive mechanisms of cobalamin-mediated methylation, and oxidative

and "organomercurial lyase"-mediated demethylation, establish domains under varying

environmental conditions. While aerobic demethylation in estuarine sediment appears to

proceed by the organomercurial lyase pathway, oxidative demethylation appears to be the

dominant pathway in anaerobic estuarine sediment and in anaerobic and aerobic

freshwater sediment (Oremland et al., 1991).

Ratios of methylation to demethylation (M/D) that are greater than one, in the

water column, demonstrate the dominance of methylation in the water column (Korthals

and Winfrey, 1987). Peak M/D ratios in surface sediment suggest that this

microenvironment may play a significant role in mercury bioavailability, while decreased

M/D ratios in deep sediment suggest that buried mercury, in the absence of bioturbation,

is rendered unavailable for biomagnification.










Bioaccumulation in Fish


Extensive studies have been carried out to identify the factors that influence

mercury accumulation in fish (Bodaly et al., 1984; Cope et al., 1990; Verta, 1990; Nilsson

and Hakanson, 1992). Mercury concentrations in fish are typically inversely related to

ambient pH, alkalinity, and primary production (Cope et al., 1990; Haines et al., 1992;

Lange et al., 1993), and are directly related to transparency (Lange et al, 1993). Fish

mercury concentrations have been shown to increase in newly flooded reservoirs (Bodaly

et al., 1984; Johnston et al., 1991). These increases result under new flooding conditions

when increased mercury methylation, a consequence of the microbial decomposition of

dead biomass, enhances the release of labile mercury from inundated soils.


Identification and Assessment of Mercury Contamination


Concerns over regional mercury contamination typically stem from discoveries of

local contamination. Publicity about mercury contamination in the Canadian province of

Saskatchewan, in the 1970s increased dramatically after the discovery of sediment

mercury contamination in the North and South Saskatchewan Rivers (Merkowsky et al

1990). Subsequent studies were feverishly implemented to examine the cause and extent

of contamination. Evans (1986) studied mercury contamination in remote lakes of South

Central Ontario and quantified an average anthropogenic mercury loading (0.79 mg m"2)

to these lakes, with a final conclusion that the anthropogenic loading of mercury to these

remote lakes came from direct atmospheric deposition from outside the catchment area.

Once direct causal relationships of environmental mercury contamination can be identified








36

in studied systems, regional control measures may be implemented to ameliorate presumed

widespread contamination (Hamdy and Post, 1985).

Many questions remain regarding the abundance, transport, and cycling of

mercury, and the long term ecological effects of mercury contamination are uncertain

(Fitzgerald and Clarkson, 1991). However, sufficient information exists regarding the

contribution of anthropogenic activities to the global mercury budget (Clarkson et al.,

1984), the human health hazard, and major factors influencing mercury bioaccumulation,

to facilitate informed regulatory and remediation decisions.


Paleolimnological Studies


Paleolimnological studies suggest widespread recent and long-term increases of

mercury accumulation in aquatic systems, resulting from changes in the global and

regional mercury cycle (De Lacerda et al., 1991, Vincente-Beckett et al., 1991; Swain et

al., 1992; Wood et al., 1992,). Decreases of mercury accumulation have resulted in

systems where mercury point sources have been eliminated (Cenci et al., 1991). Regional

sediment mercury increases have been linked to local mercury sources (Simola and

Lodenius, 1982) and to global atmospheric mercury increases (Meger, 1986; Steinnes and

Andersson, 1991) .

Sediment cores from four lakes in Minnesota were dated radiochemically after y-

assay for unsupported 21"Pb (Henning et al., 1989). The resulting mercury accumulation

rates suggested significant increases in mercury accumulation since the turn of the

century. Pre-1850 mercury accumulation rates ranged from 4 to 15 tg m' yr', and








37

present mercury accumulation rates ranged from 10 to 100 pg m2 yr". Mercury

concentrations in surface sediment were enriched 264% (80% to 450%, n=12) as

compared to deep sediment mercury concentrations.

Tolonen et al. (1988) dated sediment cores from the Baltic Sea near Oulu, Finland,

using the 210Pb dating technique. The resulting age-depth relationship corresponded very

well with "varve" dates. Mercury accumulation rates from this location ranged from

approximately 50 jg m-2 yr2 in 1920 to a peak accumulation rate (4200 pg m2 yr') in

1970, with a decline to 700 jlg m2 yr-1 after 1980. The mid-century increases were

directly related to mercury discharges from a chlorine manufacturing facility.

Mercury accumulation rate profiles for dated sediment cores from seven lakes in

Wisconsin and Minnesota suggested that recent atmospheric mercury deposition rates were

3.4 times greater than those from pre-industrial times (3.7 to 12.5 Pg m-2 yr')(Swain et

al, 1992). The resulting 2% average annual increase was compared to an estimated 1.5%

annual increase in atmospheric mercury concentrations over the North Atlantic Ocean

(Slemr and Langer, 1992) to suggest that global atmospheric increases were the primary

determinant of mercury accumulation in those aquatic systems (Swain et al. 1992).

Norwegian sediment cores from ombrotrophic peat bogs showed increased mercury

concentrations from <50 ng g-' in deep sediment (50 cm) to -190 ng g-' in surface

sediment (Steinnes and Andersson, 1991). These increases corresponded to 188% (57%

to 363%, n=l 1) enrichment of mercury in surface sediment. Forest soil cores, receiving

atmospheric mercury inputs originating from a cement factory in Japan (Fukuzaki et al.

1986), suggested 183% (43% to 318%, n=5) enrichment of mercury in surface soils

compared to deep strata (40-50 cm).












Summary


Improved technology has enabled researchers to characterize the abundance,

speciation, and transport of mercury in the environment. As a result, current research,

has examined details of the global mercury cycle with greater interpretive resolution than

was previously possible.

Approximately half of the mercury inputs to the atmosphere are derived from

human-related activities. While particulate-phase emissions of mercury may establish

regional gradients of mercury in air, water, sediment, and biota, most mercury emissions

(90 to 99%) readily enter the global atmospheric mercury cycle as elemental mercury

vapor. Photo-oxidative and particulate scavenging mechanisms facilitate the conversion

of vapor-phase mercury to a particulate form that is subsequently deposited on the earth's

surface. Mercury forms strong associations with soil and sediment matrices, however,

biotic and abiotic processes facilitate the release of small quantities (<1%) of mercury to

overlying water, primarily as aqueous monomethylmercury. Monomethylmercury readily

accumulates in biota and it's concentration biomagnifies along the food web. At this time,

the complexity of the global mercury cycle, and the remaining technological barriers,

preclude a comprehensive assessment of ecological and human-health hazards imposed

by present and historic mercury contamination. However, the current understanding of

mercury behavior in the world ecosystem provides a strong foundation to compare and

contrast mercury contamination in environmental systems.














CHAPTER 3
MATERIALS AND METHODS


Site Selection


Sampling sites were selected to encompass a spectrum of conditions of

hydroperiod, soil type, and human impact (agriculture, urbanization), in an attempt to:


1) determine natural baseline mercury content and accumulation,

2) identify human-related changes in mercury content, accumulation,
and transport, and

3) characterize associations between mercury distribution and
selected physicochemical parameters.


Samples were retrieved from sites in seven major hydrologic regions described as: Water

Conservation Areas 1, 2, and 3 (WCA), the Stormwater Treatment Areas (STA) within

the Everglades Agricultural Area (EAA), and the Everglades National Park (ENP)(Figure

3.1); the Okefenokee Swamp (OKE)(Figure 3.2); and Savannas State Reserve

(SAV)(Figure 3.3). The sampling regime was also chosen to optimize sampling of

transitional areas (i.e. agriculturally impacted vs. unimpacted). In regions with significant

water level variability, sediment cores were retrieved from the wet areas rather than

nearby dry areas. Sediment cores were collected when possible, and soil grab samples

were collected in the few cases that a sediment core could not be obtained.




























3I 0


Figure 3.1. Sample locations in the Florida Everglades


26'30'-








26'00'-








25'30'-








25'00' -


80'30'


80'00'
I


81'30'


81'00'






















820 15'00"


310 00'00" -











300 45'00" -


OKE:56
OKE:57
OKE:58


Suwannee
River


Figure 3.2. Sample locations in the Okefenokee Swamp


820 07'30"


820 22'30"


820 30'00"






















80'25'


27'30'


SAV:50


27'25'


SAV:54


SAV:53
SAV:55


27'20'--


SAV:49


SAV:48


27'15'J-


SR 707A


Figure 3.3. Sample locations in the Savannas State Reserve


80'20'


80'30'


80'15'










Field Sampling


Transport to sampling sites in the Water Conservation Areas, Everglades National

Park, and the Everglades Agricultural Area was provided by the South Florida Water

Management District using an airboat or a pontoon-equipped helicopter. The geographic

coordinates for sites accessed by helicopter were converted to latitude/longitude

coordinates from Global Positioning System (GPS) coordinates measured using on-board

equipment. Sample locations in the Savannas State Reserve (SAV) and the Okefenokee

Swamp (OKE) were accessed by foot, or by canoe. Sample coordinates for these

locations were determined using quadrangle maps (latitude/longitude).

Temperature, conductivity, and dissolved oxygen of surface waters were measured

mi situ using YSI (Yellow Springs Instruments) portable field meters and pH was

measured using a Fisher Scientific Accumet portable pH meter. Sediment cores were

obtained using thick-walled polyvinyl chloride (PVC) tubing (7.5 cm diameter, 80 to 100

cm length). Core barrels were inserted slowly into the sediment matrix to minimize

compaction. Once inserted, the top of the core barrel was capped with a large rubber

stopper. The core barrel was maneuvered from side to side and then pulled from the

substrate. The bottom of the core was then sealed with a large rubber stopper and the top

stopper was removed to fill the top of the core barrel with water to reduce any movement

of the sediment. The top rubber stopper was then replaced and both stoppers were taped

securely with duct tape. Cores were transported upright to the base camp for extrusion.

Sediment compaction averaged 27% (17-36%, n = 9).

For extrusion of the sediment, core tubes were attached to a vertical galvanized

pipe. A piston was inserted into the bottom of the core barrel. The core barrel was








44

lowered while the piston was held stationary and two centimeter sections of sediment

were removed, sequentially, from the top of the core barrel. The core extrusion was

continued until the entire core was sectioned from the surface to deeper strata. Core

sections were transferred to previously labelled Whirlpak bags. Sample bag labels

included the sample identification number, date of sampling, and the initials of personnel

involved with core extrusion. All sediment samples were stored in the dark at 40C in an

insulated chest during field operations and transported to the laboratory. Samples were

then placed in a freezer until sample analysis was initiated.


Total Mercury


Total mercury was determined using the digestion procedure described in EPA

method 7471 for the determination of mercury in soil and sediment followed by cold

vapor atomic absorption spectrophotometry (U.S.E.P.A., 1986). Sediment samples were

mixed in the Whirlpak sample bags, using an acid rinsed teflon-coated spatula, and two

grams of wet sample were transferred and weighed (to 0.0001 g) into a 10 mL plastic

beaker cup on a Mettler AE-160 analytical balance. The sample was transferred

quantitatively to an acid-rinsed 300 mL BOD bottle with a 10 mL deionized water rinse.

The digestion involved addition of 2.5 mL of concentrated nitric acid and 5 mL of

concentrated sulfuric acid. The sample was heated at 950C for two minutes, then 15 mL

of potassium permanganate (50 g L'), and 8 mL of ammonium peroxydisulfate (50 g L')

were added to the digestion mixture. The sample was then heated at 950C for one hour.

An additional 15 mL of potassium permanganate solution was added to the digestion








45

mixture if the permanganate color disappeared within fifteen minutes of the initial

addition. Upon completion of digestion, samples were cooled and decolorized by the

addition of 6 mL of hydroxylamine hydrochloride solution (120 g hydroxylamine sulfate,

and 120 g sodium chloride per liter of deionized water).

Each digested sediment sample was transferred to a plastic reaction vessel fitted

for a Perkin Elmer MHS-10 cold vapor unit. Stannous chloride solution (80 g L') was

added continuously (10 mL per minute) to the digestate in the reaction vessel. The

sample was continuously purged with high purity nitrogen gas. Elemental mercury was

evolved from the digestate and swept with the nitrogen purge-gas into an open ended

quartz tube (1 cm diameter) with a 16 cm cell path length. The mercury was quantified

by cold vapor atomic absorption spectrophotometry using a Perkin Elmer model 5000

Atomic Absorption Spectrophotometer (X=253.6 nm, SBW=0.7 nm) with a mercury

hollow cathode lamp (1=6 mA). Light absorption was measured as peak height. The

standard calibration curve working range (0 to 50 ng Hg) gave an absorbance range from

0.003 to 0.035 absorbance units. The detection limit for mercury analysis was 10 ng g'.


Percent Solids/Bulk Density


Percent solids were determined by weighing known volumes of wet sediment m

aluminum weighing dishes. Wet sediment was transferred into a 25 cm3 glass syringe that

was modified to function as a piston chamber. The empty dish was weighed, a known

volume of wet sediment was transferred to the dish and weighed. The sample was dried

in an oven for 24 hours at 1040C, removed, and placed in a desiccator for approximately








46

1 hour. The dried sample was then re-weighed. The wet and dry sample weights were

corrected for the weight of the empty dish. Percent solids were then calculated as the

percent of dry mass to total wet mass. Bulk density was determined from the same

aliquot of wet sediment. The dry bulk density of the sample was calculated as the dry

sediment mass per 10 cm3 sample volume (g cm").


Radionuclide Analysis


To calculate age/depth relationships in sediment cores, the activity of unsupported

2t"Pb was estimated by determining total and supported 210Pb activity. Supported 210Pb

results from, and is maintained by, radioactive decay of 226Ra (half-life 1622 years) in the

sediments. Unsupported 210Pb is formed by decay of 226Ra to 222Rn (half-life 3.8 days).

which escapes to the atmosphere, decays to 21'Pb, and is deposited to sediment via

precipitation. Subtracting supported 2"Pb from the total measured activity of 21"Pb in

sediment samples yields the unsupported 2o"Pb activity, that will decrease with depth in

the sediments because of radioactive decay. The age of a sediment layer may then be

calculated from its activity of unsupported 21oPb. Because the half-life of 21oPb is only

22.3 years, this dating technique is restricted to about a 150-year time span. Activity of

'"Cs serves as an independent age marker because it first appeared in the atmosphere

during nuclear bomb testing around 1960.

Activities of 2o"Pb and '"Cs were measured by direct y-assay using two intrinsic-

germanium well-detectors (Princeton Gamma Tech). This type of detector counts over

a large range of y-energies and can be used for simultaneous measurement of supported








47

and unsupported 21Pb (Gaggeler et al., 1976), as well as '7Cs which may be used as an

additional age-marker (Ritchie et al., 1973). Lead shielding (10.1 cm thick) was used to

reduce natural background radiation at the germanium detector. Samples for radionuclide

analysis were dried at 950C for 24 hours, pulverized by mortar and pestle, weighed, and

placed in small low-density polypropylene tubes (capacity 4 mL). The volume of the

samples and standard were matched to ensure the same counting efficiencies for both.

Core sections were combined (up to 2 cm) to obtain an adequate sample volume. Sample

tubes were sealed with plastic cement and left for a minimum of 14 days to equilibrate

radon (R22Rn) with radium (226Ra).

Counting times varied from 7 to 26 hours depending on sample weight; small

samples needed longer counting times to minimize uncertainty. For each region of

interest, counts were corrected for Compton scattering by subtracting the below-the-peak

area from the total counts. This area was determined by a linear fit through three channel

contents (e.g. counts) on either side of the region of interest.

Blanks were counted for every two samples to determine background from ambient

radiation. Standards (Department of Energy, New Brunswick Laboratories: U-Th

standards) were run with the same frequency to track efficiency (counts y-') and to

calculate a Z26Ra conversion factor (pCi counts"' s-1). Sample spectra were analyzed for

activity m the 46.5 keV (210Pb) and 662 keV ('37Cs) peaks. Activities at 295 keV (214pb),

352 keV (14pb), and 609 keV (214Bi) representing uranium series peaks were used to

compute supported 210Pb abundance.








48

Calculation of 21"Pb dates followed the constant rate of supply (CRS) model

(Goldberg, 1963) which is able to quantify changing sediment accumulation rates. This

model appears applicable to Florida aquatic systems, particularly because 2"OPb residuals

match both the known atmospheric flux of this isotope as well as the residuals of nearby

cores (Binford and Brenner, 1986; Gottgens, 1992). These residuals are defined as the

total inventory of unsupported 21"Pb (pCi cm') in the core from the surface to the depth

at which its activity has decayed to background levels. Such a constant rate of 210Pb

fallout is likely, due to the high efficiency at which 21oPb is scavenged from the

atmosphere and from the water column by wet precipitation or particulate matter

(Turekian et al., 1977; Robbins, 1978). This provides evidence favoring the assumption

of the CRS dating model that an increase in the rate of delivery of bulk sediments will

not supply more 2"OPb. Finally, a constant rate of 21oPb fallout will result in different

unsupported 2"oPb activities at the sediment-water interface in core locations with differing

rates of net sediment accumulation. This has been confirmed by paleolimnological

investigations in aquatic systems throughout Florida (Binford and Brenner, 1986).

Uncertainty analyses were based on both the random variation of counting errors

associated with radioactive decay and the nature of the CRS model. Errors controlled by

external forces such as inaccuracies of stratigraphic sampling and determination of bulk

density were not considered.

Radiation emitted in nuclear decay is subject to statistical fluctuation. This

unavoidable source of uncertainty is often a predominant source of imprecision (Knoll,

1979). Because the recorded counts in nuclear counting experiments follow a Poisson








49

distribution, the predicted standard deviations were estimated as the square root of the

mean number of counts. The amount of 210Pb (total, supported, and unsupported) and

'"Cs was expressed as activity (pCi g1') one standard deviation (i.e. 68.3% confidence

limits), which is standard practice in expressing uncertainty in nuclear measurements

(Wang et al., 1975; Binford, 1990). Counting errors in the calculation of net isotope-

activities were propagated using first-order analysis.

Monte Carlo simulation (Palisade Corp., 1990) was used to estimate error

associated with the calculation of age and sedimentation rate following the CRS model.

The probability density function for simulated 210Pb activities was approximated by a

normal distribution with the mean equal to the measured activity and a range equal to the

counting error.


Carbon


Total Carbon


Total carbon was analyzed using a Coulometer (Coulometrics, Inc., Model 5011)

combined with a Total Carbon Combustion Apparatus (Coulometrics, Inc., Model 5020).

Total carbon measurements were made by weighing approximately 5 mg of air dried

sediment into a platinum boat. The platinum boat, containing the dried sample, was

placed in the entry port of a preheated (9500C) furnace. Contaminant CO2 was swept

through the furnace to the attached coulometric cell. The coulometer solution was titrated

coulometrically to eliminate contaminant interference. The platinum boat, containing the

sample, was then moved from the entry port into the furnace. Carbon dioxide, evolved








50

from the sample, was swept into the coulometric cell. The resultant pH change induced

a color change in the coulometric solution. The solution was then titrated coulometrically

to the initial pH and color. The analysis quantified in units of micrograms carbon and

percent of total carbon was calculated.


Inorganic and Organic Carbon


Inorganic carbon was analyzed using the coulometric procedure, described above,

coupled with a Carbonate Carbon Apparatus (Coulometrics, Inc., Model 5030). Dry

sediment (10 to 20 mg) was transferred to a porcelain boat, placed in a glass tube, and

attached to the Carbonate Carbon Apparatus. The glass tube was placed on a heating

element and 3 mL of perchloric acid (2 N) was introduced to the sample. Carbon

dioxide, evolved from the sample, was swept into the coulometric solution and titrated

(Huffman, 1977; Lee and Macalady, 1989). Organic carbon was determined as the

difference between the total and inorganic carbon content.


Additional Trace Metals


Analyses for cadmium (Cd), chromium (Cr), copper (Cu), iron (Fe), nickel (Ni),

lead (Pb), and zinc (Zn), were performed on 0.5 to 1 gram dried sediment aliquots using

the digestion procedure described in EPA Method 3050 (U.S.E.P.A., 1986). Ten mL of

1:1 nitric acid (HNO3) were added to the sediment in a beaker and covered with a watch

glass. The mixture was heated to 950C and refluxed for 10 to 15 minutes without boiling.

The sample was cooled and 5 mL of concentrated HNO3 was added. The watch glass








51

was replaced and the solution was allowed to reflux for 30 minutes. The last step was

repeated to ensure complete oxidation. Covered with the watch glass, the solution was

then concentrated by evaporation to 5 mL without boiling. Two mL of deionized (DI)

water and 3 mL of 30% hydrogen peroxide (HzO,) were added to the cooled solution.

The beaker was covered with the watch glass and was warmed on the hot plate to initiate

the peroxide reaction. Hydrogen peroxide was added in 1 mL aliquots, with warming,

until effervescence became minimal or until the general sample appearance was

unchanged. Not more than 10 mL of 30% HzO2 were added to minimize acid dilution

and digestate volume. Next, 5 mL of concentrated hydrochloric acid (HC1) and 10 mL

of DI water were added to the solution, covered and returned to the hot plate to reflux for

an additional 15 minutes without boiling. After cooling, the solution was filtered through

Whatman No. 41 filter paper to remove particulates. The filtrate was diluted to 100 mL

with DI water. The acid concentration was 5.0% (v/v) HCI and 5.0% (v/v) HNO3 for the

diluted solution.

Metals were quantified, by flame atomic absorption spectrophotometry (FAAS),

using a Perkin Elmer model 5000 Atomic Absorption Spectrophotometer, with appropriate

hollow cathode lamps and an air/acetylene flame. The following instrument settings

(Table 3.1) and detection limits (Table 3.2) were determined.
















Table 3.1. Instrument settings for metal analyses using a Perkin
Elmer model 5000 Atomic Absorption Spectrophotometer

Element Wavelength Bandwidth
(k, nm) (nm)

Cr 357.9 0.7
Pb 283.3 0.7
Ni 232.0 0.2
Cd 228.8 0.7
Zn 213.9 0.7
Cu 324.7 0.7
Fe 248.3 0.2


Table 3.2. Detection limits for metals determination using a Perkin
Elmer Model 5000 Atomic Absorption Spectrophotometer (Flame
Atomizer)


Analyte


Detection Limit
(mg/Kg dry weight)














CHAPTER 4
RESULTS AND DISCUSSION


Water Quality


Sample site coordinates (latitude/longitude) and associated water quality parameters

(depth, temperature, conductivity, dissolved oxygen, and pH) are presented in Table 4.1.

These data demonstrate the variability of water quality and quantity throughout the

Everglades region. Water depth at Everglades sites (ENP, WCA1, WCA2, WCA3, and

STA) ranged from 0 to 0.6 meters. Water depth at the Savannas sites (SAV) ranged from

0.1 to 1.4 meters. Okefenokee sites (OKE) were covered by a floating Sphagnum spp.

mat (approximately 0.5 m thick) with approximately 0.5 meter of underlying water.

Conductivity of overlying water ranged from 49 to 37000 gmhos cm' for the

Everglades. Average conductivities for WCA1, WCA2 and WCA3 were 257, 1257, and

625 pmhos cm', respectively, while those for the ENP were regionally variable, with a

measured range from 465 to 37000 Lmhos cm-'. Water at the periphery of WCAI is

supplied to some degree by agricultural runoff and exhibits higher conductivities (230 to

850 pmhos cm-), while most of the water in the center of WCA1 is derived from

precipitation (49 to 98 pmhos cm-'). The conductivities of water at OKE and SAV

sample sites also indicate the predominance of precipitation to the regional hydrology

(OKE, 42 to 121 pmhos cm-', SAV, 54 to 74 Vnmhos cm-1, respectively).













Table 4.1. Water quality data associated with wetland sediment sample sites

Sample ID# Latitude Longitude Depth Temp. [DO] Cond. pH
(m) (deg. C) (mg/L) (umhos/cm)

ENP:01 254303 804311 0.10 24.5 6.4 465 8.0
ENP:02 254121 803809 0.10 26.0 2.8 780 7.6
ENP:03 253101 803802 0.00 N/A N/A N/A N/A
ENP:04 253647 804129 0.10 26.0 5.9 500 7.9
ENP:05A 252004 804450 0.00 N/A N/A N/A N/A
ENP :05B 252004 804450 0.00 N/A N/A N/A N/A
ENP:05C 252004 804450 0.00 N/A N/A N/A N/A
ENP:05D 252004 804450 0.00 N/A N/A N/A N/A
ENP:05E 252004 804450 0.00 N/A N/A N/A N/A
ENP :05F 252004 804450 0.00 N/A N/A N/A N/A
ENP:06 251457 803608 trace 30.0 1.0 8200 7.7
ENP:07 251705 803805 0.05 27.0 2.0 500 7.6
ENP :08A 252754 805114 0.00 N/A N/A N/A N/A
ENP :08B 252754 805114 0.00 N/A N/A N/A N/A
ENP:09 253625 811014 0.03 21.8 1.0 37000 7.2
ENP :10 253201 810011 trace 25.0 4.3 23000 7.1
ENP:11 253119 804741 0.10 26.0 3.5 1000 7.4
ENP: 12 253632 805632 0.13 26.0 5.0 625 7.4
WCA3:13 255024 804944 0.45 27.0 6.8 325 7.8
WCA3:14 254959 804156 0.45 26.0 5.6 405 7.4
WCA3:15 254953 803305 0.15 28.5 8.5 700 7.8
WCA3:16 255707 802905 0.15 25.5 2.4 750 7.5
WCA3:17 255702 804151 0.45 26.0 4.2 480 7.4
WCA3:18 260400 803805 0.30 21.5 1.5 460 7.4
WCA3:19 260401 804804 0.30 21.5 3.8 500 7.2
WCA3:20 261802 804754 0.00 N/A N/A N/A N/A
WCA3:21 261014 804457 0.05 ND ND ND ND
WCA3:22 260914 804201 0.30 23.0 2.4 650 7.5
WCA3:23 261756 803652 0.10 25.0 7.9 900 6.0
WCA3:24 261002 803302 0.15 25.0 3.0 800 7.2
WCA2:25 261041 802156 0.15 25.0 2.0 900 7.1
WCA2:26A 261800 802056 0.15 25.0 6.3 1350 7.3
WCA2:26B 261800 802056 0.15 25.0 6.3 1350 7.3
WCA2:27 262555 802652 0.10 15.5 7.2 1200 7.5
WCA2:28 261901 802658 0.15 17.0 2.7 1325 7.4
WCA2:29 262149 802058 0.05 16.0 1.9 1350 7.8
WCA2:30 262034 802030 0.30 17.0 1.8 1250 7.3
WCA2:31 261954 802105 0.15 17.5 1.5 1425 7.3
WCA3:32 260147 802855 0.45 19.5 2.3 800 7.3
WCA3:33 255923 803053 0.60 19.8 2.7 700 7.3

N/A corresponds to sites with no overlying water
ND corresponds to data not determined













Table 4.1. (cont'd)

Sample ID# Latitude Longitude Depth Temp. [DO] Cond. pH
(m) (deg. C) (mg/L) (umhos/cm)

WCA3:34 255739 803219 0.45 20.5 2.5 650 7.3
WCA1:35 264005 802141 0.30 26.5 0.5 850 6.7
WCA1:36 263449 802047 0.20 29.5 2.7 90 7.0
WCA1:37 262924 801939 0.55 30.5 2.1 82 6.6
WCA1:38 262806 802441 0.55 30.0 1.8 442 6.7
WCA1:39 263200 802447 0.10 30.0 3.0 230 7.1
WCAI:40 262258 801657 0.25 31.0 1.6 49 7.5
WCA1:41 262719 801452 0.30 30.5 1.3 98 7.4
WCAl:42 263304 801543 0.35 34.2 1.8 218 7.6
STA:43 263919 802510 0.30 35.0 1.2 530 7.5
STA:44 263736 802526 0.00 N/A N/A N/A N/A
STA:45 263854 802440 0.10 32.0 2.3 500 8.5
STA :46 263842 802537 0.30 34.0 0.5 560 7.5
STA :47 263927 802436 0.00 N/A N/A N/A N/A
SAV :48 271630 801500 1.4 20.2 6.2 114 ND
SAV:49 271645 801530 1.1 22.4 7.6 121 ND
SAV:50 272115 801830 0.00 N/A N/A N/A N/A
SAV:53 272000 801750 1.0 18.0 ND 42 ND
SAV :54 272015 801730 1.0 18.0 ND 72 ND
SAV :55 271945 801700 0.1 17.8 ND 78 ND
OKE :56 304235 821000 0.5 17.0 2.4 74 ND
OKE:57 304235 821000 0.5 17.2 5.3 68 ND
OKE :58 304235 821000 0.5 19.0 7.2 54 ND

N/A corresponds to sites with no overlying water
ND corresponds to data not determined








56

High conductivity water in Everglades National Park (ENP) is related to estuanne

mixing while high conductivity water at Water Conservation Area (WCA) sampling sites

is an indicator of hydrologic inputs from the Everglades Agricultural Area (EAA). Low

conductivity of surface water in the center of WCA 1 has been used to demonstrate that

the primary hydrologic source is from precipitation (Richardson et al., 1990; SFWMD,

1992).


Sediment Geochronology


Results of paleolimnological analyses may be presented in units of concentration

or as rates of accumulation. Concentration, expressed as a relative measure of sediment

composition (e.g. mg g-'), is the conventional way of expressing sediment stratigraphy

(Shapiro et al., 1971; Pennington, 1973; Griffiths and Edmondson, 1975). Such data,

however, are vulnerable to variations in sedimentation of other components in the profile.

These variations may result in dilution of the target analyte. This problem can be

eliminated by using ratios of components in the sediment matrix, or by calculating

accumulation rates. The latter are normalized to time thus avoiding the problem of co-

variance among sedimentary components.

Compilations of all data for sediment cores retrieved from the Everglades,

Okefenokee Swamp, and Savannas State Reserve appear m the Appendix. Blank cells in

the tables of the Appendix occurred so that all data could be presented in a consistent

tabular format. The tables include total mercury, solids, bulk density, total and organic

carbon, cadmium, copper, chromium, iron, lead, nickel, and zinc. These data are

summarized in Table 4.2. The Appendix also includes aspects of sediment geochronology














'O. 0 m m(o rso




r zi oock r ) >



\ov t kr n W VIi m


-o









0



0












a)




-S 0

0




a)
M'






















0


>U
au












Va)




















>
4m
0t
o
^
0 ?
^^














e ^



























d


:4

























0
oo




16
m h


mp m '0 -b It







mi 0 'n 6m c T o




m. m so o~ o


o


("4


t- On 0 ro ( oo 00
D 0\1 N Iv 0N O \t


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5 4 V
(1) WMM


o-
.0 00
<'


c00 -N -C Cm


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E


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00









.
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0 3


en Cu
CO
o m

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* *








58

based on 21"Pb dating (i.e. sediment accumulation rate, mercury accumulation rate, and

age/depth relationships). Total mercury, percent solids, bulk density, and water quality

were measured for all sample sites, while sediment geochronology, total and organic

carbon, and additional metals were determined for selected samples.


Sediment Dating Acceptance Criteria


Radiochemical techniques are routinely used to date lake sediment profiles. Few

attempts have been made, however, to apply these methods in wetland sediment.

Diagenesis in wetland deposits is poorly understood and the correlation between depth and

time-of-deposit may be affected by compaction, decomposition, and vertical migration of

the element for which sedimentation rates are computed. Compaction may be accounted

for by calculating material deposition in units of mass (grams) rather than depth (cm) over

time. Decomposition of organic matter may increase the concentration of the analyte of

concern (e.g. 21"Pb, mercury, and others). Because of the nature of the CRS model,

however, core sections with such concentrated 210pb (C) will have proportionally lower

deposition rates for bulk sediment (r) and, thus, for the analyte of concern (its

concentration multiplied by the bulk sedimentation rate). This follows from the CRS-

calculation for sedimentation rate according to


r k (1)
C

where

r = bulk sediment accumulation rate (g.cm2.yr1)
A = the residual 2"OPb beneath the sediment horizon of interest (pCi cm2)
k = 2"OPb radioactive decay constant (yrf'), and
C = unsupported 2"OPb activity in the sediment horizon of interest (pCi g-).








59

The potential for temporal variability in the depositional environment in a wetland

may limit the resolution of a core's age/depth profile. Burning of dry, organic wetland

soil, for example, may cause a loss of 21OPb from the profile to the atmosphere (fly-ash).

This reduces the cumulative residual 21"Pb, i.e. the amount of this isotope (pCi cm-2) in

the core from the surface to the depth at which its concentration has decayed to

background level. Such reduction makes age/depth determinations less accurate.

Confidence in the dated profiles is enhanced when cumulative residual 210Pb corresponds

among cores from the same area despite differences in sediment accumulation rates.

The cores analyzed from this 5600 km2 Everglades area showed an average 210Pb

residual of 15.5 pCi cm-2 (sd = 3.5, n = 20). The range of residuals corresponded to 21OPb

fallout rates between 0.33 and 0.67 pCi cm'2 y-', which was well within the normal range

of 21oPb fallout of 0.2-0.9 pCi cm"2 y-1 (Appleby and Oldfield, 1983). Seven cores with

fallout rates outside this range were excluded from the analysis (Figure 4.1). These

profiles may have been disturbed over time by removal or addition of material (producing

a lower or a higher cumulative residual 21oPb, respectively).

Additional support for age/depth relationships may come from matching peak-'7Cs

activity in the profile with a 2"Pb-determined age of 1963 (Krishnaswami and Lal, 1978).

These peaks (or the onset of '"Cs activity in the absence of a distinct peak) occurred in

the profiles (n = 18) at an average 21"Pb-determmed age of 1962, although the range of

age-values was considerable (1942-1978) (Figure 4.2). This may suggest some post-

depositional mobility of '"Cs (up or down in the core). An additional two cores

(WCA 1:36; WCA3:17) were excluded from consideration because their assigned dates to

peak '"Cs activities fell outside this range (Figure 4.2).












210Pb Fallout Rate (pCi cm-2yr-)
0.33 0.67


c O V (0 MO O C 0
|o a o -- o Nc


Cumulative Residual 210pb (pCi/cm2)

Figure 4.1. Cumulative residual unsupported 210Pb (pCi cm"2) for all cores
analyzed radiochemically. Cores with fallout rates outside the range 0.33-0.67 pCi
cm2 yr-1 were not included in the computation of material accumulation rates.
Fallout of 2'"Pb is the product of the 2 oPb residual and the radioactive decay
constant for 210Pb. Core identifications are placed within the bars.


2000

1980-

1960-

1940-

1920-

1900-

1880


.- 0 .
... ....................... ............. ....


0*!




0


WCA 1


10 0
.......o o








0


WCA 2 WCA 3


Figure 4.2. 2'oPb determined age of the core section with peak activity of '"Cs
(*) for dated sediment cores from Water Conservation Areas 1.2. and 3.
Everglades National Park; and Savannas State Reserve. Open data points (0)
represent profiles in which the onset of '"Cs activity was used as a marker
honzon m the absence of a distinct '"Cs peak.










Sediment Mercury Geochronology


The geochronology of cores that satisfied the above described sediment dating

acceptance criteria are presented in Figures 4.3 through 4.20. Recent (post-1985) and

historic (approximately 1900) average sediment accumulation rates for each sample region

are given in Table 4.3. The mercury accumulation rate is calculated as the product of the

sediment accumulation rate and the total mercury concentration at each depth interval of

the sediment profile. Turn of the 20th century (ca. 1900) and recent (post-1985) mercury

accumulation rates for dated cores are averaged by sample region (Table 4.4).








Table 4.3. Recent and historic average sediment accumulation rates in cores retrieved
from Water Conservation Areas 1, 2, and 3, Everglades National Park, and Savannas State
Reserve. Numbers in parentheses indicate the range of values found.

Sample Number of Average Sediment Accumulation Rate (g cm2 y-')
Region Cores 1900 Post-1985

WCA1 5 0.018 (0.009-0.030) 0.047 (0.016-0.099)
WCA2 3 0.021 (0.011-0.030) 0.042 (0.031-0.064)
WCA3 3 0.015 (0.009-0.023) 0.069 (0.029-0.143)
ENP 5 0.033 (0.015-0.054) 0.060 (0.044-0.075)
SAV 2 0.019 (0.016-0.023) 0.027 (0.024-0.030)








62

Table 4.4. Recent and historic average mercury accumulation rates in cores retrieved
from Water Conservation Areas 1, 2, and 3, Everglades National Park, and Savannas State
Reserve. Numbers in parentheses indicate the range of values found.

Average Mercury Accumulation
Sample Number of Rate (Vg m"2 y-1) Ratio "
Region Cores 1900 Post-1985 Post-1985/1900

WCA1 5 14 (5-29) 79 (45-141) 7.8 (1.6-13.3)
WCA2 3 8 (4-12) 59 (35-95) 8.7 (3.9-13.9)
WCA3 3 10 (7-11) 39 (28-55) 4.0 (3.0-4.9)
ENP 5 14 (2-28) 40 (23-57) 5.9 (1.6-19.1)
SAV 2 10 (10) 34 (31-37) 3.4 (3.0-3.8)

1) The ratio given represents the average of the ratios for the different cores in each area,
rather than the ratio of the average for each area.


Mercury accumulation rates around the turn of the 20th century ranged from 2 to

29 pg m2 y-' for all cores, with apparent increasing trends beginning mid-century (1930-

1960). Post-1985 mercury accumulation rates were an average of 6.3 (1.6-19.1) times

higher than 1900 rates. Temporal changes in average mercury accumulation rates

progressed geographically from a 5.9 and 4.0 times increase in ENP and WCA1 (post-

1985/1900), to a 7.8 and a 8.7 times increase in WCAI and WCA2 (Table 4.4). Average

mercury accumulation rates for SAV cores increased 3.4 times (post-1985/1900). The

trend of larger ratios to the north (WCA 1,WCA2) with smaller ratios to the south (WCA3,

ENP) suggests at least three possible explanations:

1) some northern source of mercury in overland sheetflow;

2) non-uniform atmospheric deposition of mercury with more deposition
in northern regions;

3) non-uniform, post-depositional mobility of mercury in soils
(i.e. varying retention of mercury in different soil types) with
mercury retention decreasing spatially from WCA1 south to
ENP.




















Water Conservatic
Unsup 210Pb (pCi/g) 137Cs
0 3 6 9 12 0 2
0,

-10

-20 --

/ -30 ,

-40

-50


Depth in Core (cm)
-45 -30 -15 0
2000

1970

1940

1910

1880

1850

Totol Hg (ng/g)
0 25 50 75 100
2000

1970

1940 i

1910

1880

1850


0

-10

E -2C
o
0

_ -3C

-40

-50C


Figure 4.3. Sediment paleostratigraphy for Water Conservation Area 1--Core 01


1n Area 1 Core 01

(pCi/g) Bulk Density (g/cm3)
4 6 8 0.00 0.06 0 2 0.18
0

-10

-- -20

-30

-40

-50

Sed.Rt (g cm-2y-1)
0.00 0.05 0.10
2000 ,0
0.24
1970

1940

1910

1880

1850

Tot.Hg Acc.Rt. (ug m-2y-1)
0 25 50 75 100
2000
181
1970

1940

1910

1880

1850


-50



















Water Conservation Area
Unsupp.210pb (pCi/g) 137Cs (pCi/g)
0 3 6 9 12 15 0 3 6 9 12 15
0 0- .

-5 0 -5

-10 / -10

0 -15 1 -15

-20 -20
a.
S-25 -25

-30 -30

Depth in Core (cm) Sed.Rt.
-30 -20 -10 0 0.00 C
2000 -- 2000

S1970 1970
.o
o 1940 1940
a-

S1910 1910

S1880 1880

1850 1850


Total Hg (ng/g)
0 50 100150200
2000
2QOOi --------..

1970

1940

1910

1880

1850
--- Detection limit


1 Core 35
Bulk Density (g/cm3)
0.0 0.1 0.2
0

-5 /

-10 \

-15 \

-20

-25

-30

(g cm-2-1)
.03 0.06 0.09


Tot.Hg Acc.Rt. (ug m-2y-1)
0 20 40 60 80
2000

1970

1940

1910

1880

1850


Figure 4.4. Sediment paleostratigraphy for Water Conservation Area 1--Core 35



















Water
Unsupp.210pb (pCi/g)
0 10 20
0


-10 .


-20
!

-30 )


-40

Depth i
-35
2000

c 1970

o 1940

1910
0
1880

1850


Total
0 4(
2000

1970 /

1940

1910

1880

1850 --


Conservation Area
137Cs (pCi/g)
0 3 6 9
0

-10


-20 /


Core 37
Bulk Density (g/cm-1
0.00 0.05 0.10
0


10


20
'\


-30 4 -30


-40 -' -40

i Core (cm) Sed.Rt. (g cm-2y-1)
-20 -5 0.00 0.03 0.06
S2000

1970

1940

1910

1880

-1850

Hg (ng/g) Tot.Hg Acc.Rt. (ug m-2y-1)
00 800 0 100 200
2000

1970

1940

1910

1880

S1850


--- Detection limit


Figure 4.5. Sediment paleostratigraphy for Water Conservation Area 1--Core 37


I



















Water Conservation Area 1 Core 38
Unsupp 2100b tpCi/g) S-Cs (.Ci/g) Sulk Densivt (q/cm-)
0 5 10 15 20 0 2 5 0.0 0. .1 0.2
0 0 0 .

-5 -5 -5

-10 / -10 -10

-15 -15 -15

-20 -20 -20

-25 -25 -25

Depth in Core (cm) Sed.Rt. (g cm-2- 1)
-20-15-10-5 0 0.00 0.03 0.06
2000 2000

c 1970 1970
o U
o 1940 1940
0.
o
7 1910 1910

S1880 1880

1850 1 1850

Total Hg (ng/g) Tot.Hg Acc.Rt (ug m-2y-1)
0 150 300 0 100 200
2000 2000

c 1970 1970

0 1940 1940
0 .

o 1910 1910
0
1880 1880

1850 I 1850
--- Detection limit


Figure 4.6. Sediment paleostratigraphy for Water Conservation Area I--Core 38


















Water Conservation Area 1 Core 40
Unsupp.210Pb (pCi/g) 137Cs (pC./g) Bulk Density (g/cm3)
0 4 8 12 16 0 2 4 6 8 0.00 0.05 0.10 0.15
0 0 O
/ K. ,
-5 -5 -5

S-10 '-10 /*/ -10 \
o 0
- -15 / -15 -15

-20 -20 -20

-25 -25 -- -25

Depth in Core (cm) Sed.Rt. (g cm-2y-1)
-20-15-10-5 0 0.00 0.02 0.04
2000 2000

c 1970 1970
o0
o 1940 1940
)
1910 1910
0
S1880 1880

1850 1850

Total Hg (ng/g) Tot.Hg Acc.Rt. (ug m-2y-1)
0 200 400 0 50 100
2000 2000

c 1970 / 1970

o 1940 1940

S1910 1910

S1880 1880

1850 -- 1850
--- Detection limit


Figure 4.7. Sediment paleostratigraphy for Water Conservation Area 1--Core 40






















Unsupp C
0 6 1





S/"

/'


Figure 4.8. Sediment paleostratigraphy for Water Conservation Area 2--Core 25


0

-5
E
u
E -10
0

-15
r-

o -20

-25


Water Conservation Area 2 Core 25
OPb (pCi,'g) 137Cs (pCi/g) Bulk Density (g/cm3)
2 1 24 0 2 4 6 0.0 0.1 0.2
0 0






-15 -15 -

-20 -20

S -25 -25

Depth in Core (cm) Sed.Rt (g cm-2y-1,
-25 -15 -5 0.00 0.02 0.04
2000 2000

1970 1970

194C 1940

191C 1910

1880 / 1880

1850 1850

Total Hg (ng/g) Tot.Hg Acc.Rt. (ug m-2y- )
0 200 400 600 0 50 100150200
2000 2000

1970 1970

1940 1940

1910 1910

1880 1880

1850 1850
--- Detection limit




















Water Conservatior
Unsupp.210Pb (pCi/g) 1 37Cs
0 3 6 9 12 0 1


0

-5 -"
I
-10


I -15
-20

-25

Depth in Core (cm)
-20-15-10-5 0
2000

c 1970
.o
o 1940
a.

S1910

- 1880

1850


Total Hg (ng/g)
0 50 100150200
2000

1970

1940 "

1910

1880 I

1850
--- Detection limit


C
0
E


o
r-

a.
X -2C

-25


Figure 4.9. Sediment paleostratigraphy for Water Conservation Area 2--Core 26


j


Area 2 Core 26
(pCi/g) Bulk Density (g/cm3)
2 3 0.0 0.1 0.2 0.3
O

-5 (

-10

-15 /

-20

-25

Sed.Rt. (g cm-2y-1)
0.00 0.04 0.08
2000

1970

1940

1910

1880

1850

Tot.Hg. Acc.Rt. (ug m-2y-1)
0 20 40 60
2000

1970

1940

1910

1880

1850




















Water Conservation Area 2
Unsup.210pb (pCi/g) 137Cs (pCi/g)
0 2 4 6 8 0 2 4 6 8 10


Depth in Core (cm)
-35 -25 -15 -5
2000

1970

1940

1910

1880

1850

Total Hg (ng/g)
0 50 100150200
2000
I -

1970 1-

1940

1910 \

1880 i

1850 '
--- Detection limit


- Core 29
Bulk Density (g/cm3)
0.0 0.1 0.2 0.3
0 -------


0


E -10
l-i
0
u -20
C
-E
" -30


-40


Figure 4.10. Sediment paleostratigraphy for Water Conservation Area 2--Core 29


-40

Sed.Rt. (g cm-2y-1)
0.00 0.03 0.06
2000

1970

1940

1910

1880

1850

Tot.Hg Acc.Rt. (ug m-2y-1
0 20 40 60


1970

1940

1910

1880

1850



















Water Conservatior

Unsupp.210pb (pCi/g) 137Cs
0 4 8 12 16 0 2


Depth in Core (cm)
-20-15-10 -5 0
2000

1970

1940

1910

1880

1850

Totol Hg (ng/g)
0 50 100150200
2000
/"
1970 \

1940 .

1910 :

1880

1850
--- Detection limit


n Area 3 Core 13

(pCi/g) Bulk Density (g/cm3)
4 6 0.0 0.1 0.2 0.3
0


-5 /


-10

\
-15


-20

Sed.Rt. (g cm-2y-1)
0.00 0.03 0.06


1970

1940

1910

1880

1850

Tot.Hg Acc Rt. (ug m-2y-1)
0 20 40 60
2000

1970

1940

1910

1880

1850


Figure 4.11. Sediment paleostratigraphy for Water Conservation Area 3--Core 13


-5


-10


-15


-20

















Water

Unsupp.210Pb (pCi/g)
0 5 10 15 20


-5 )


Conservation Area 3 -

137Cs (pCi/g)
0 2 4 6 8


* -- u -
1 /

4/ -15


-20

Depth in Core (cm)
-20-15-10-5 0
2000

0 1970
0
o 1940
a
S1910
0
1 1880

1850

Total Hg (ng/g)
0 60 120180240
2000

c 1970 -

0 1940 (
4) \
1910 /

1880 /

1850
--- Detection limit


Core 15

Bulk Density (g/cm3)
0.0 0.1 0.2
0


-5


-10

/
-15 \


1' '- -20 L

Sed.Rt. (g cm-2y-1)
0.00 0.02 0.04
2000

1970

1940

1910

1880

1850

Tot.Hg Acc.Rt (ug m-2y-1)
0 20 40 60
2000

1970

1940

1910

1880

1850


Figure 4.12. Sediment paleostratigraphy for Water Conservation Area 3--Core 15


-10


-15


-20



















Water Conservation Area 3 Core 19
Unsupp.21OPb (pCi/g) 137Cs (pCi/g) Bulk Density (g/cm3)
0 3 6 9 12 0 3 6 9 12 0.0 0.1 0.2
0 0 0


E -5 -5 -5


o -10 -10 -10


) -15 -15 / -15


-20 -20 -0n


Depth in Core (cm)
-20-15-10 -5 0
2000

c 1970
0

o 1940
0.

S1910
0 '
S1880
/
1850

Total Hg (ng/g)
0 50 100150200
2000

c 1970
0
in
o 1940

S1910

I 1880

1850
--- Detection limit


Sed.Rt. (g cm-2y-1)
0.0 0.1 0.2 0.3
2000

1970

1940

1910

1880

1850

Tot.Hg Acc.Rt. (ug m-2 -1)
0 30 60 90
2000

1970

1940

1910

1880

1850


Figure 4.13. Sediment paleostratigraphy for Water Conservation Area 3--Core 19



















Taylor Slough: Core 1(.--), Core 2 (.......... )
Unsup.210pb (pCi/g) 37Cs (pCi/g) Bulk Density (9~ cm3)
0 3 6 9 12 0.0 0.5 1.0 '.5 0.0 0.3 0.6 .9
0-- 0 0--

? -5 -5 -5 *-

T -10 -10 -10 .
o .
S-15 -15-15 -15

3 -20 *-20 -20

-25 -25 -25

Depth in Core (cm) Sed.Rt. (g cm-2y-1)
-25 -15 -5 0.0 0.1 0.2
2000 2000

c 1970 1970
0
o 1940 1940

1910 / 1910

1880 1880-

1850- 1850 -,

Total Hg (ng/g) Tot.Hg Acc.Rt. (ug m-2y-1)
0 25 50 75 100 0 20 40 60 80
2000 2000

c 1970 1970
o

S1940 1940

1910 1910

>- 1880 1880

1850 --L 1850
--- Detection limit



Figure 4.14. Sediment paleostratigraphy for Everglades National Park--Taylor Slough



















Everglades Natic
Unsuo 210Pb (pCi/g) 137Cs
0 2 4 6 8 0.0 0.5 1

S-5



/ -10


-15


-20

Depth in Core (cm)
-20-15-10 -5 0
2000

c 1970

o 1940

o 1910

1880

1850


Totol Hg (ng/g)
0 30 60 90 120
2000

1970 \

1940 K

1910

1880 *

1850 Detecn
--- Detection limit


onai Park Core 7
(pCi/q) Bulk Density (g/cm3)
.0 .5 2.0 0.0 0.2 0.4
0


\...-5",


-10


-15 \


'-20

Sed.Rt. (g cm-2y 1)
0.00 0.05 0.10
2000

1970 -

1940 /

1910

1880

1850

Tot.Hg Acc.Rt. (ug m-2y1)
0 30 60 90
2000

1970

1940

1910

1880

1850


Figure 4.15. Sediment paleostratigraphy for Everglades National Park--Core 7


-20


,I


















E\erglades Natior

Unsupp.210Pb (pCi/g) 137Cs
0 3 6 9 12 0.0 0.5 1
0 0
/
.5 -5 /




5 -15


0 -20


Depth in Core (cm)
-20-15-10-5 0
2000

1970

1940 i

1910

1880 L

1850

Total Hg (ng/g)
0 50 100150200
2000

1970

1940 /

1910

1880

1850 -
--- Detection limit


al Park -

(pCi/g)
0 1.5 2.0


Core 09

Bulk Density (g/cm3)
0.0 0.1 0.2
0

/
-5 [


-1n /


\
-15 /


-20

Sed.Rt. (g cm-2y 1)
0.00 0.03 0.06
2000

1970

1940

1910

1880

1850

Tot.Hg Acc.Rt. (ug m-2y-1)
0 20 40 60
2000

1970

1940

1910

1880

1850


Figure 4.16. Sediment paleostratigraphy for Everglades National Park--Core 9


-1


-1


-2

















Everalaoes Nationoi Park

Unsupp.210Pb (pCi/g) 137Cs (pCi/g)
0 3 2 0 2 4
0

-5

*/ -10 \ *


/ -15


-20

Depth in Core (cm) Sed
-20-15-10-5 0 0.00
2000 2000

c 1970 / 1970

o 1940[ 1940
0
o 1910 1910

w 1880 / 1880

1850' 1850

Total Hg (ng/g) Tot.Hg A
0 50 100150200 0
2000 2000

c 1970 1970
.0 /
-i
0 1940 1940
0
S1910 1910

1880 1880

1850 --- 1850
--- Detection limit


- Core 1 1

Bulk Density (g/cm3)
0.0 0.1 0.2 0.3
0


-5 /

-10


-15


-20 -

.Rt. (g cm-2y-)
0.03 0.06













icc.Rt. (ug m-2y-1)
20 40 60


I


Figure 4.17. Sediment paleostratigraphy for Everglades National Park--Core 11


-5

-10


-20




















Everglades National Park -
unsuc= 210Pb (pC,/g) '37Cs (pC/g)
0 4 8 12 0 2 4 6 8
0 \ i o0 I


-20 L'

Depth in Core (cm)
-20-15-10 -5 0
2000

1970

1940

1910

1880

1850


Total Hg (ng/g)


2000

1970

1940

1910

1880

1850


Core 12
Bulk Density (g/cm l
0.0 0.1 0.2 0.3
a,---


-5


-10


-15


-20

Sed.Rt (g cm-2y1)
0.00 0.03 0.06
2000

1970

1940

1910

1880

1850

Tot.Hg Acc.Rt. (ug m-2y~ '
0 30 60 90
2000

1970

1940

1910

1880

1850


--- Detection limit


Figure 4.18. Sediment paleostratigraphy for Everglades National Park--Core 12


-5 )

-15
-10'


-15


-20



















Savannas
Unsupp. 210Pb (pCi/g)
0 5 10 15 20
0


-5

/


5
5


O ,


State Reserve
13Cs (pCi/g)
0 2 4 6 8
-5




-10


-15


-20 '

Depth in Core (cm)
-15 -10 -5 0
2000

1970

1940

1910

1880

1850


- Core 48
Bulk Density (g/cm3)
0.0 0.2 0.4
0


-5 \


-10 ,


-15


-20 'L

Sed.Rt. (g cm-2y-1)
0.00 0.05 0.10
2000

1970

1940

1910

1880

1850


Totol Hg (ng/g)
0 50 100 150
2000

1970 .

1940
I'
1910

1880

1850
--- Detection limit


Tot.Hg Acc.Rt (ug m-2y-)
0 2C 40 60
2000

1970

1940

1910

1880 i

1850


Figure 4.19. Sediment paleostratigraphy for Savannas State Reserve--Core 48


-1


-1


-2




















Savannas
Unsupp. 210Pb (pCi/g)
0 5 10 15 20
0 --


-5


-10
I

-15 I


-20


State Reserve
137Cs (pCi/g)
0 2 4 6 8
0


-5 ."


-10 F

/
-15
rL.


- Core 49
Bulk Density (g/cm3)
0.0 0.2 0.4
0


-5


-10


-15


epth in Core (cm)
-10 -5 0


Totol Hg (ng/g)
0 50 100 150


('.








i


----- -20

Sed.Rt. (g cm-2y-1)
0.00 0.05 0.10
2000

1970

1940

1910

1880

1850

Tot.Hg Acc.Rt. (ug m-2y-1)
0 20 40 60
2000

1970

1940

1910

1880

1850


--- Detection limit


Figure 4.20. Sediment paleostratigraphy for Savannas State Reserve--Core 49


2000

1970

1940

1910

1880

1850




2000

1970

1940

1910

1880

1850


- Cl ,


J


--








81

Agricultural runoff is delivered to the Everglades from agricultural land to the

north (Reddy et al., 1991). Agricultural practices in the Everglades Agricultural Area

have increased erosion and oxidation of organic soils (Blake, 1980). Historically,

mercurial fungicides were used in this region to enhance agricultural production. Many

agricultural practices (i.e. repeated drying and flooding, and the application of mercurial

fungicides), and the resulting erosion and oxidation of soils, can facilitate the transport

of mercury (Lodenius et al., 1987) from agricultural land to surrounding areas. Although

there is no basis, from existing data, to quantify the relative contribution of agricultural

practices in the EAA to mercury accumulation in the Everglades system, previous studies

have demonstrated the deleterious effects of agricultural runoff on the Everglades wetland

system (Horvath et al., 1972; Richardson et al., 1990; Reddy et al., 1991). The northern

portions of the Everglades (WCA1 and WCA2) likely receive atmospheric and drainage

inputs of mercury from the Everglades Agricultural Area (EAA) via mill production and

the burning of crop material (121 Kg y-'; 1981 to 1990)(KBN Engineering and Applied

Science, 1992) and the irrigation of agricultural runoff waters (Richardson et al., 1990;

Reddy et al., 1991).

A Florida mercury emissions survey identified four primary anthropogenic sources

of mercury in 1990, including MSW (municipal solid waste) combustion (14.6%), medical

waste incineration (14.0%), paint application (11.1%), and electricity production (10.7%)

(KBN Engineering and Applied Sciences, Inc., 1992). Natural processes contributed

38.9% of the total 1990 mercury emissions. The survey did not identify the relative

contributions of these sources to mercury deposition in the state. Globally, studies have








82

identified significant discharges of mercury from manufacturing and incineration activities

(Fukuzaki et al., 1986), and regional gradients of mercury accumulation from point source

emissions have been identified (Nater and Grigal, 1992). Since there is little direct

quantitative information regarding mercury deposition in Florida, at this time, spatial

differences in mercury emissions must be used as a surrogate measure of mercury

deposition.

Palm Beach, Broward, and Dade counties follow the southeast coastline of Florida

such that Palm Beach county resides east of WCA1, Broward county resides east of

WCA2 and northern WCA3, while Dade county resides east of southern WCA3 and ENP.

The estimated total 1990 mercury emissions for Palm Beach, Broward, and Dade counties

were 1512, 1995, and 4614 Kg y', respectively (KBN Engineering and Applied Sciences,

1992).

To predict non-uniform mercury deposition resulting from local anthropogenic

activities, one would expect to find the greatest mercury enrichment to occur in southern

regions of the Florida Everglades (i.e. ENP and WCA3). The reverse trend is suggested

by the sediment data (Table 4.4). Alternatively, it must be recognized that mean

estimates of mercury accumulation rates for the hydrologic basins (WCA's and ENP)

resulted from soil cores with considerable between-site variability. Further, the central

regions of Water Conservation Areas 1 and 2 are closer m proximity to the Atlantic coast

of peninsular Florida (20 and 30 km, respectively), than those of Water Conservation Area

3 and Everglades National Park (45 and 60 km, respectively).

Variable retention of mercury between organic soils and marl sediments may

influence the ambient concentrations in these substrates. The three Water Conservation









83

Areas have organic-rich soils, with a total organic carbon content (g g-') between 40 and

50%. Sediment in ENP represents an array of marl and organic deposits, with a total

organic carbon content of 10-20%. If mercury retention were variable between organic

and marl deposits, and mercury inputs were uniformly distributed, then post-depositional

mercury migration would alter the apparent mercury accumulation rates in the mineral

sediment of ENP (Barrow and Cox, 1992a, 1992b).

Mercury retention by organic and mineral substrates, can be compared by

examining average mercury accumulation rates among the Everglades regions (WCAs and

ENP). For this purpose, let us assume:


1) atmospheric mercury deposition serves as the primary mercury
input to WCA3 and ENP soil,

2) mercury deposition is uniform over WCA3 and ENP soil,
and

3) mercury retention by organic soil exceeds mercury
retention by mineral sediment.



One would predict that mercury accumulation rates in ENP sediment would be less than

contemporaneous rates in WCA3 soil. Assumptions 1 and 2 likely pertain to pre-

development (1900) mercury accumulation rates in ENP and the WCAs. The average

1900 mercury accumulation rate for ENP cores (14 tg m"2 yr') is similar to average

mercury accumulation rates for WCA1, WCA2, and WCA3 cores (14, 8, 10 lg m"- yr-,

respectively). Further, average post-1985 accumulation rates are similar for ENP and

WCA3 cores (40 and 39 Lg m-2 yr-, respectively). Enhanced mobility of mercury, that








84

may occur in marl sediment (Barrow and Cox, 1992a, 1992b), is not demonstrated by the

dated cores examined in this study.

The data suggest that mercury retention is not influenced by variability in soil

composition. Further, increases in accumulation rate ratios for Everglades regions moving

from south (ENP) to north (WCA1)(Table 4.4) suggest a regional factors) that results in

more pronounced mercury accumulation mi the northern Everglades (WCA 1 and WCA2)

as compared to WCA3 and ENP.

Mercury accumulation rates increase gradually since the turn of the century and

increase more distinctly by mid-century (1930-1960). Twelve of the eighteen mercury

accumulation profiles exhibit dramatic increases beginning in the 1970's and 1980's

(Figures 4.3-4.20). Some sites show increased mercury accumulation during the last two

decades with constant sediment accumulation rates (WCA1:37, WCA2:26, ENP:07) or,

alternatively, with uniform mercury concentration (WCAI:01, WCA2:29, WCA3:19,

ENP:09, ENP: 11). The covariance of sediment component inputs demonstrated in these

cores illustrates the limitations of characterizing mercury deposition using mercury

concentration profiles alone.

Gradual post-1900 increases in mercury accumulation rates match trends found in

other systems. These increases are probably related to global atmospheric increases

resulting from European and American industrialization since the turn of the 20th century.

Mid-century increases in accumulation rate are likely related in part to regional

urbanization and agriculture in south Florida (Blake, 1980).

Urban development along the southeast coast of peninsular Florida expanded

dramatically since the 1940's (Blake, 1980). Municipal solid waste incineration began in








85

1951 on Florida's southeast coast. Between 1951 and 1972, the construction of eleven

incinerator facilities was completed in Dade and Broward counties. All of these facilities

were shut down by 1979 as a consequence of their inefficient emission controls. The

cumulative mercury emissions from waste incineration in southeast Florida increased from

955 Kg y-' in 1951 to a maximum of 1870 Kg y-' in 1973 with a decrease to 0 Kg y"' in

1979 (KBN Engineering and Applied Sciences, 1992). Mercury emissions from

modernized incineration facilities, constructed since 1983, have increased due to increased

facility throughput (tons per year)(Table 4.5).






Table 4.5. Mercury emission estimates for municipal solid waste incineration in southeast
Florida (Palm Beach, Broward, and Dade counties)(KBN Engineering and Applied
Sciences, Inc., 1992).

Year Low Average High
(Kg y-') (Kg y-') (Kg y-')

1982 0 0 0
1983 504 575 671
1984 457 522 610
1985 388 443 518
1986 272 310 362
1987 389 445 519
1988 342 391 456
1989 447 511 597
1990 700 834 1105
1991 1225 1471 1776








86

Statewide mercury emissions m Florida, from the electric utilities industry, were

estimated to have increased 51%, from 2,062 Kg y-' between 1981-82 to 3,111 Kg y'

between 1989-90 (KBN Engineering and Applied Science, 1992). Southeast Florida

mercury emissions, in 1990, averaged 79 (8-203) Kg y"' from electricity production, 835

and 1,820 Kg y-' from MSW and medical waste incineration, respectively. Despite

improved emission controls, increasing mercury emissions since 1980, result from the

demands (utilities and waste control) imposed by rapidly developing regions in the state.


Error Analysis of Sediment Dating


Errors associated with the statistical fluctuations of nuclear decay and with the

application of this uncertainty in the CRS dating model were determined for three

different sites (WCAI:01, ENP:11, and SAV:49). Uncertainty for all other sites was

assumed to be of similar magnitude.

"Error bars" (one standard deviation on either side of the data point) for activity

of radiochemicals are shown in Figures 4.21-4.23. Because the predicted standard

deviation for random processes, such as gamma disintegrations, equals the square root of

the mean count, samples with a high count have a small standard deviation (as percent

of that mean). Standard deviations generally ranged from 3-6% of the mean for the

higher activity deposits to 6-30% for deeper core sections. Errors in the activity of

unsupported 2"oPb were larger (generally 3-12% and 12-54%, respectively), because they

are computed as the difference of two uncertain activities.

Dating uncertainty increased with age of the sediment (Figures 4.24-4.26). Monte

Carlo simulations (Palisade Corp., 1990) were used to calculate 500 different 2"OPb










Activity (pCi/g)


2 4 6


Water Conservation Area


1 Core 1


Figure 4.21. Error associated with radionuclide determinations for WCA1:01.


-5


-15


-25


-35


-45


E
o
(D
0
0

C-

C.
Q-
0


-5


-15


-25


-35


-45









Activity (pCi/g)
6 9


-5


-10


-15


-20

0


-5


-10


-15


-20


Everglades National Park -


Core 1 1


Figure 4.22. Error associated with radionuclide determinations for ENP: 11.











Activity (pCi/g)

10 15


2 4 6


Savannas


State


Reserve


- Core 49


Figure 4.23. Error associated with radionuclide determinations for SAV:49.


20


E

0)
L-
0
O
C
Ic
(-
(-
Q_
a
0-


-5




-10


-15

0


E
o
/)
0
0

C
c

a
Q)


-5




-10


-15










Depth in Core (cm)


-45
2000 r-


-35


Sedimentation Rate (g cm-2y 1)


0.0


1990


1980


1970


1960


Water Conservation Area


1 Core 1


Figure 4.24. Dating uncertainty associated with WCAI:01.


-25


-15


-5


1970

1940

1910

1880

1850

1820


0.1


0.2


0.3


_ r I 1


I 1










Depth in Core (cm)


-20
2000 r-


1970

1940

1910

1880

1850

1820


0.00
2000

1970

1940

1910

1880

1850


1820


Sedimentation Rate (g cm 2y 1)


0.02


0.04


0.06


Everglades National Park -


Core 1 1


Figure 4.25. Dating uncertainty associated with ENP:11.


-15


-10


-5


0.08









Depth in Core (cm)


-10


-5


0.00
2000

1970

1940

1910

1880

1850


Sedimentation Rate (g cm-2y1)


0.02


0.04


0.06


Savannas State


Reserve


- Core 49


Figure 4.26. Dating uncertainty associated with SAV:49.


2000

1970

1940

1910

1880

1850




Full Text
SPATIAL AND TEMPORAL DISTRIBUTION OF MERCURY AND OTHER
METALS IN FLORIDA EVERGLADES AND SAVANNAS MARSH SOILS
By
BRIAN EUGENE ROOD
A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL
OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT
OF THE REQUIREMENTS FOR THE DEGREE OF
DOCTOR OF PHILOSOPHY
UNIVERSITY OF FLORIDA
1993

Copyright 1993
by
BRIAN EUGENE ROOD

This dissertation is dedicated to my parents, F. Eugene and Roberta Rood, the best
teachers I've known. I also dedicate this dissertation to Professors Edward S. Deevev, Jr.
(deceased) and Peter H. Rich for prompting me to recognize the power of the imagination

ACKNOWLEDGMENTS
I wish to thank Dr. Joseph Delfino, for his supervision of my research, and my
committee members, Drs Ronnie Best, Emmett Bolch, Donald Graetz, and Frank Nordlie
for their critical review of this dissertation. Dr. Claire Schelske kindly reviewed my
dissertation and attended my oral defense. I gratefully acknowledge laboratory assistance
by William Beddow, Candace Biggerstaff, Celia Earle, Becky Fierle, Ingrid Forbes,
Lizanne Garcia, Manuel Llahues, Kathleen Newell, Margaret Olson, Brandon Selle,
Marcia Sommer, and Melissa Voss. Special thanks go to Richard Pfeuffer, Liberta Scotto,
and Bob Przekop for their assistance with planning and implementation of field sampling,
and to Curtis Watkms, Dr Thomas Atkeson, Thomas Swihart (Flonda Department of
Environmental Protection) and Larry Fink (South Florida Water Management District),
who served as project officers for the funding agencies. This research was funded by
grants from the Florida Department of Environmental Protection, South Florida Water
Management District, and the United States Geological Survey. I am greatly mdebted to
my friend. Dr Johan F Gottgens, for his assistance, generosity, and candor, throughout
my doctoral studies, and to Brian Cutchens and Lola Wilcox for their treasured friendship.
IV

TABLE OF CONTENTS
ACKNOWLEDGMENTS iv
ABSTRACT vii
CHAPTER 1
INTRODUCTION I
Background 1
Present Study 3
CHAPTER 2
REVIEW OF LITERATURE 8
Overview 8
Human-Related Activities 9
Mercury Issues m Florida 11
Available Technology for Mercury Research 12
Global Mercury Cycle 17
Global and Regional Interactions 19
Mercury m the Atmosphere 20
Mercury m Water 21
Mercury m Sediment 26
Mercury m Biota 29
Environmental Factors and Bioaccumulation 31
Mercury Transformations m Aquatic Systems 31
Bioaccumulation m Fish 35
Identification and Assessment of Mercury Contamination 35
Paleolimnological Studies 36
Summary 38
CHAPTER 3
MATERIALS AND METHODS 39
Site Selection 39
Field Sampling 43
Total Mercury 44
Percent Solids/Bulk Density 45
Radionuclide Analysis 46
v

Carbon 49
Total Carbon 49
Inorganic and Organic Carbon 50
Additional Trace Metals 50
CHAPTER 4
RESULTS AND DISCUSSION 53
Water Quality 53
Sediment Geochronology 56
Sediment Datmg Acceptance Criteria 58
Error Analysis of Sediment Datmg 86
Sediment Mercury Concentrations 94
Comparison of Recent and Histone Mercury Concentrations 94
Post-Depositional Mobility of Mercury 99
Error Analysis of Mercury Determinations 102
Spatial Distnbution of Mercury in the Everglades 102
Relationships Between Mercury Concentration and Selected Water
and Sediment Parameters 106
Supplementary Sediment Metals Concentrations 108
CHAPTER 5
SUMMARY, CONCLUSIONS, AND RECOMMENDATIONS 120
Summary 120
Conclusions 121
Recommendations 122
REFERENCE LIST 124
APPENDIX
FLORIDA WETLAND SOIL CHEMISTRY DATABASE 142
BIOGRAPHICAL SKETCH 180
vi

Abstract of Dissertation Presented to the Graduate School
of the University of Florida in Partial Fulfillment of the
Requirements for the Degree of Doctor of Philosophy
SPATIAL AND TEMPORAL DISTRIBUTION OF MERCURY AND OTHER
METALS IN FLORIDA EVERGLADES AND SAVANNAS MARSH SOILS
By
Brian Eugene Rood
December 1993
Chairperson: Joseph J Delfino
Major Department: Environmental Engmeermg Sciences
Elevated mercury concentrations were identified previously m freshwater fish in
the Everglades, Savannas State Reserve, and receiving waters of the Okefenokee Swamp
The goals of this research were to 1) determine historic baselme concentrations of
mercury m Florida wetland soils, 2) determme post-development changes m sedimentary
mercury accumulation, and 3) identify the spatial distribution of mercury throughout the
Florida Everglades. Sixty soil cores were analyzed for total mercury. Selected cores
were analyzed for carbon and trace metals, and were chronologically analyzed after
radionuclide analysis for 2I0Pb and l37Cs.
The average mercury concentration m surface sediment (0-4 cm) of 121 ng g1
(n=51, 17-411 ng g'1) was 2.5 times (0.2-10.6, n=51) higher than corresponding deep
sediment (11-17 cm) concentrations. The largest increases were measured m Water
vii

Conservation Areas 1 and 2 (3.7 times higher for both) of the Florida Everglades, while
Okefenokee Swamp sediment showed the smallest relative increase (1.4). Because
concentration data are vulnerable to temporal variations in bulk sediment accumulation
rate, the interpretive problem of co-vanance was avoided by determining mercury
accumulation rates after radionuclide dating. Post-1985 mercury accumulation rates
averaged 53 pg m2 y'1 (23-141 pg m'2 y’1) corresponding to a 6.4 (1.6-19.1, n= 18) times
rate increase smce the year 1900. The largest rate increases occurred in WCA1 and
WCA2 cores (7.8 and 8.7 times higher, respectively), while the Savannas State Reserve
cores showed the smallest rate increase (3.4). Mercury accumulation rates increase
starting about 1940, due perhaps to mid-century alteration of the hydrologic structure of
the Everglades, and to mcreased regional agricultural and urban development. There is
presently insufficient information regarding regional inputs to quantify any direct causal
relationship between mercury accumulation rate increases and regional human activities
However, apparent nonuniform accumulation of mercury m the Everglades hydrologic
basins, coupled with mcreased accumulation rates of other trace metals, indicate some
atmospheric contribution of mercury from regional anthropogenic activities. The findings
are similar to trends reported for lakes m Minnesota, Wisconsm, and Sweden This
agreement is significant, perhaps indicating a global process that leads to similar
accumulation rates over widely varying geographic regions. This research provides the
first data on mercury accumulation in subtropical wetland systems and demonstrates the
feasibility of radiochemical datmg of wetland cores.
viii

CHAPTER 1
INTRODUCTION
Background
A statewide survey of mercury concentrations in sportfish was implemented after
preliminary indications of mercury contamination appeared in Florida freshwater fish
(Hand and Friedemann, 1990). The survey revealed mercury concentrations in fish m the
Everglades (Water Conservation Areas 1 and 2) and Savannas State Reserve that exceeded
acceptable levels for human consumption. Numerous lakes, rivers, and wetlands yielded
fish with mercury concentrations sufficient to warrant limited-consumption advisories.
Tins survey identified the magnitude of fish mercury contamination m the state.
However, the survey did not address issues regardmg the origin, transport, and availability
of mercury in these habitats.
Concurrent studies of wildlife suggested that mercury is transported through the
food web of the Florida Everglades and that the viability of the endangered Florida
panther has been dimmished due to mercury bioaccumulation (Roelke et ak, 1991). The
risk to other animal populations from mercury biomagmfication has been suggested and
the potential for perturbations of ecosystem structure and function have been examined
(Jurczyk, 1993).

~)
Recent increases of mercury accumulation rate have been reported for north
temperate lake systems in Sweden. Wisconsin, and Minnesota (Meger, 1986; Wiener et
ah, 1990; Lmdqvist et ah, 1991; Swam et ah, 1992). In some cases, atmospheric
deposition could account for mcreased mercury accumulation rates in recent sediment
(Meger, 1986). Some studies have linked a 1.5% annual increase of atmospheric mercury
concentrations (1977-1990) (Slemr and Langer, 1992) to an estimated 2% increase in
mercury deposition rates in Wisconsm and Mmnesota (Swam et ah, 1992). These studies
suggested that mercury deposited on the surface and watershed of remote lake systems
originated from regional or global atmospheric sources (Swam et al., 1992).
It is estimated that about 95 percent of atmospheric mercury occurs in the gaseous
elemental form, with an atmospheric residence time of 0.7-2.0 years (Nater and Grigal.
1992). Approximately 5 percent of atmospheric mercury is associated with particulates
(Fitzgerald et ah, 1991) which can readily be deposited as dryfall or scavenged from the
atmosphere during ram episodes (Fitzgerald, 1986). Anthropogenic emissions of
elemental mercury enter the global atmospheric cycle and may be distributed far from
their source; however, particulate phase mercury from emission sources may establish
regional concentration gradients in nearby soils (Nater and Grigal, 1992).
Urban emissions of mercury from industry (i.e. cement production), medical and
municipal waste mcmeration, fossil fuel combustion, m addition to agricultural emissions
(i.e. burning of crop material, volatilization of mercurial fungicides), may contribute to
atmospheric emissions and eventual deposition of mercury (Crockett and Kinnison, 1977;
Fukuzaki et ah, 1986; Sengar et ah, 1989; KBN Engmeermg and Applied Sciences, Inc..
1992).

3
Perturbations of the natural hydroperiod may facilitate the release of historically
accumulated mercury because of changes in the physical properties of soils. Oxidation
and deep crackmg of dried agricultural land may release both naturally and
anth ropo gem cal ly derived mercury and facilitate its transport to wetland soils in runoff
(Del Debbio, 1991). Because flooded peat soils readily accumulate trace metals by
adsorption and sulfide precipitation they serve as a primary smk for mobilized mercury
(Lodenius et ah, 1987; Norton et ah, 1990).
Present Study
The Florida Everglades, Savannas State Reserve, and the receivmg waters of the
Okefenokee Swamp (Suwannee and Santa Fe rivers) exhibited elevated fish mercury
concentrations (Hand and Friedemann, 1990). These aquatic systems are unique Florida
habitats, and there is concern that mercury contamination poses a serious ecological and
human health hazard. This study examines mercury abundance and distribution in
sediment from the Everglades, Savannas Marsh, and Okefenokee Swamp wetland systems
(Figure 1.1).
The Everglades is "perhaps the most recognized wetland m the world, its notoriety
derived from the wealth of its biotic heritage as well as the magnitude of factors that
threaten its resources" (Gunderson and Loftus, 1993, p. 1). It is a dynamic subtropical
aquatic system (5600 km2), subject to hydrologic variability, fire, and human related
activities (Blake, 1980). The Everglades are considered oligotrophic, based on dominant
plant communities and ambient nutrient concentrations, and are characterized by peat soils

4
Figure 1.1. Geographic distribution of wetland study sites.

5
to the north and marl sediment to the south. Sawgrass (Cladium íamaicense) marshes
dommate large expanses of this system. The region is spotted with intermittent wet
prairies, tree islands and shallow ponds. Durmg the past century, extensive draining of
this wetland for agriculture and diversion of water to coastal urban centers has altered its
natural hydroperiod. Regions south of Lake Okeechobee were drained for agriculture, and
canal systems were constructed to control water movement. Presently, some areas of the
remaining wetland are subject to prolonged dry periods while other locations are subject
to extended periods of inundation (SFWMD, 1992).
The Okefenokee Swamp, in southeastern Georgia and northern Florida, is the
second largest wetland m the United States (1750 km2). The flat, sandy watershed of the
Okefenokee Swamp is small (1200 km2) and siltation is negligible (Casagrande and
Erchull, 1976). As a result, precipitation serves as the predominant hydrologic mput and
filling of the wetland basm is minimal. The Okefenokee consists of an "array of diverse
habitats" includmg lakes, wet prairies with floating peat mats (Sphagnum spp), and
Taxodium spp. swamps that are mtegrated hydrologically to form one unit ecosystem.
This swamp has organic-rich soils underlam by a pure white quartz sand (Casagrande and
Erchull, 1976). The relatively pristine condition of the Okefenokee Swamp permits it to
"serve as a control for comparison with other ecosystems that contmue to be heavily
influenced by human activities" (Rykiel, 1984, p. 374).
The Savannas State Reserve is a dynamic, linear wetland system (20 km x 2 km)
just west of the Indian River Ridge m Florida's St. Lucie and Martin counties. It is a
strip of marshlands, ponds, lakes, and islands, perched -4 m above mean sea level, and

6
characterized by rich inundated muck soils overlymg relict sand dune on hardpan. The
marsh is dominated by broomsedge (Andropogon virgmicus). water lily (Nymphaea
odorata), and spatterdock (Nuphar luteum) while the surrounding watershed is a pme
(Pmus elliottii) and saw palmetto (Serranoa repens) habitat (Jurgens, 1981). This region
is considered to be "highly susceptible to damage by pollution or over-enrichment of its
water" (Davis, 1990, p. 4) due to its size and to encroaching development
Previous studies have identified atmospheric mercury deposition as a primary
vector leading to mercury accumulation m aquatic systems (Meger, 1986). Recent
increases in mercury accumulation in aquatic systems have been attributed to global
(Swam et ah, 1992) and regional (Sengar et al, 1989; Nater and Gngal, 1992) mcreases
of atmospheric mercury emissions, largely attributed to a variety of anthropogenic
activities (KBN Engmeermg and Applied Sciences, Inc., 1992). Further, elevated fish
mercury concentrations have been attributed to mcreased mercury inputs from human-
related activities (Bodaly et ah, 1984; Hakanson et ah, 1990a, 1990b). The following
study arose from concern that mcreased mercury inputs, of global or regional origin, were
causmg elevated mercury concentrations m fish.
I hypothesize that the sediment record of these subtropical wetlands will concur
with previous indications, identified in other aquatic environments, of increased mercurv
accumulation since the turn of the century. Tins hypothesis necessitates a characterization
of the feasibility of radiochemical datmg m the study wetlands. In addition, identification
of spatial variations of mercury content throughout the Everglades is essential to
characterize the relative impact of regional activities on mercury accumulation m that
system.

7
This study of Florida wetland soils was initiated in 1991 to: 1) determine the
spatial distribution of mercury throughout the Everglades, Okefenokee Swamp, and
Savannas Marsh systems, 2) identify histone baselme concentrations of mercury m Florida
wetland soils, 3) identify post-development changes m sedimentary mercury accumulation,
4) identify mercury-organic associations in wetland soils, and, 5) provide mformation to
serve as a basis for informed planning and implementation of future research and
management activities. An evaluation of spatial and temporal changes in sedimentary
mercury is necessary to elucidate the factors governing mercury accumulation and
distribution m these wetland systems.

CHAPTER 2
REVIEW OF LITERATURE
Overview
Mercury is present in air, water, soil/sediment, and biota, and is unique among the
metals with its ability to exist in the gas, liquid, and solid phases (Clarkson et al, 1984;
Moore and Ramamoorthy, 1984). The abundance of mercury in the environment is
determined by the inputs supplied by both natural and anthropogenic processes
(Fitzgerald, 1986; Mitra, 1986). Natural processes, such as volcamsm and degassing from
the land and ocean, supply a baselme mercury contribution to the atmosphere and to
water. That mercury is eventually transported to, and accumulated m, soil, sediment, and
biota. Mercury inputs to the environment may undergo numerous transformations that are
determined by physico-chemical interactions of mercury under varying environmental
conditions, including those of pH, temperature, oxidation-reduction potential, soil type,
and hvdrology (Moore and Ramamoorthy, 1984; Lodemus et ah, 1987; Del Debbio, 1991;
Barrow and Cox, 1992b). Anthropogenic activities may increase mercury mputs to the
environment or they may elicit the transport or transformation of ambient mercury
Presently, anthropogenic mercury inputs comprise approximately one-half of the
mercury entering the world ecosystem (Fitzgerald and Clarkson, 1991). There are
indications that atmospheric mercury concentrations are steadily increasing (Slemr and
8

9
Langer, 1992) and increased rates of mercury accumulation to aquatic systems have been
demonstrated in the sediment record (Meger, 1986; Norton et al, 1990; De Lacerda et
ah, 1991; Swam et ah, 1992). Elevated mercury concentrations m fish pose a human
health hazard and mercury biomagnification through the food web provides evidence of
the ecological stresses imposed by this trace metal (Cardeilhac et ah, 1981; Hand and
Friedemann, 1990; Roelke et ah, 1991; Heaton-Jones, 1992; Jurczyk, 1993). Analytical
technology has been challenged by the unique problems (i.e. low concentration, low vapor
pressure, analytical contammation, speciation, toxicity, and biomagnification) associated
with environmental mercury research (Fitzgerald, 1986; Schroeder, 1989; Douglas, 1991).
Human-Related Activities
Mercury has been used widely: 1) for the production of electrical devices, 2) as
a catalyst for the chlor-alkali industry, 3) as a fungicide/algicide in paint products,
paper/pulp manufacture, and agriculture, and 4) as a component m the manufacture of
instruments (i.e. thermometers), dental preparations, and pharmaceuticals (Mitra, 1986;
Nnagu, 1990; KBN Engineering and Applied Sciences, Inc., 1992). Anthropogenic
releases of mercury to the environment are related to activities including the bummg of
fossil fuels (Crockett and Kmnison, 1977; Sengar et ak, 1989; Lodemus, 1990), the
incineration of municipal solid and medical waste (Collins and Cole, 1990; Volland, 1991;
KBN Engmeermg and Applied Sciences, Inc., 1992), the production of electricity
(Lmdberg, 1980), wastewater discharge (Morel et ah, 1975), agricultural practices
(Simons, 1991; Patrick et ah, 1992), mmmg (Pfeiffer et af, 1991), and chlor-alkali and

10
cement manufacture (Fukuzaki et af, 1986; Mitra, 1986). Further, land development and
hydrologic manipulation facilitates the release of naturally derived mercury from deep
("old") sediment (Horvath et af, 1972; Simóla and Lodenius, 1982). Collins and Cole
(1990) outlmed a mass balance of mercury discharges to the environment (Table 2.1).
Table 2.1. Human-related discharges of mercury to the
environment (Kg yr'1).
Source
1973
1988
Industry
Chemical manufacture
307,709
18,043
Petroleum refining
36,943
227
Smeltmg
65,743
0
Electronics manufacture
185,394
1,000
Utilities
Coal burning
40,625
73,483
Natural gas
27,393
N/A
Incinerators
16,829
40,234
There is a rich, centunes-old, history of the contribution that mercury has played
in society (Fitzgerald, 1986; Mitra, 1986). Environmental mercury contamination was
initially identified m response to human tragedy, such as mass poisonmg and death (i.e.
Mmimata disease)(D'Itn, 1991), bom of the careless use of mercury and it's haphazard

disposal (Horvath et af, 1972; Hamdy and Post, 1985; Collins and Cole, 1990).
Subsequent observations of the deleterious effect of mercury (Hakanson et af, 1990a,
1990b; Scheuhammer, 1991a, 1991b) on the world ecosystem clearly identified the need
to establish stringent guidelines for the use of mercury-containing compounds in industry
and agriculture (Revis et af, 1990; Ingersoll, 1991). However, population growth and
development pressures have created new avenues by which society may contribute to the
mercury budget of the world ecosystem through the burning of fossil fuels, medical and
municipal solid waste incineration, and electricity production (Horvath et af, 1972;
Albnnck and Mitchell, 1979; Collins and Cole, 1990).
Mercury Issues in Florida
Widespread mercury contamination was identified in Florida after the discovery
of elevated mercury concentrations m fish throughout the state (Hand and Friedemann,
1990). The death of an endangered Florida panther was attributed to mercury toxicosis
(Roelke et af, 1991), and mercury accumulation was cited as a potential cause for
dramatic declines of wading bird populations (Jurczyk, 1993). A mercury emissions
survey identified municipal solid waste (MSW) and medical waste mcmeration, the
electric utility industry, and pamt application as the primary anthropogenic sources of
atmospheric mercury emissions in Florida (KBN Engineering and Applied Sciences, Lnc.,
1992), and agricultural practices have been identified as potential release mechanisms for
naturally and anthropogemcally derived mercury reserves m rich organic soils m the state
(Simons, 1991).

12
Available Technology for Mercury Research
Mercury concentrations in water and air are much less than those found in soil,
sediment, plant tissue and animal tissue (Schroeder, 1989). To understand and evaluate
the environmental impacts of mercury contamination, and the cyclmg of mercury m the
ecosystem, analysts have been faced with a stiff technical challenge (Douglas, 1991).
Ambient mercury concentrations in water and air often fall near, or below, the limits of
detection provided by many traditional analytical techniques (Bloom, 1989; LeBihan and
Cabon, 1990). Contamination during sampling, storage, and analysis of such samples may
exceed actual ambient mercury concentrations (Fitzgerald and Watras, 1989). Mercury
transport, bioaccumulation, and toxicity in the environment often depends on
environmental conditions and mercury speciation (Cope et ah, 1990; Farrell et ah, 1990;
Lodenius, 1990; Verta, 1990; Johnston et ah, 1991; Nilsson and Hakanson, 1992). As a
consequence, the constraints imposed by environmental mercury studies challenge
researchers to optimize the available analytical technology to provide a suitable basis with
which to characterize mercury abundance and transformation m the environment.
The sample matrix and mercury content must be considered when selectmg an
analytical procedure to determine mercury in environmental samples. The selected
technique must be sufficiently sensitive to quantify the anticipated mercury concentration
and must demonstrate robustness when challenged by matrix interferences inherent to
particular sample types (air, water, soil, biota). Separation and speciation issues
associated with a given sample matrix must also be addressed (Schroeder. 1989).
Cold vapor atomic absorption spectrophotometry (CVAAS) has been the standard
analytical method for mercury determmations in environmental and biological samples

13
(Winter et af, 1977; Perry et af, 1978). Numerous modifications have been developed
to decrease sampling time, to mcrease analytical sensitivity (Freimann and Schmidt. 1982;
Mateo et ah, 1990; Munaf et ah, 1990a; Welz et ah, 1992), and to facilitate mercury
speciation (Schroeder, 1989; Munaf et af, 1990b; Rapsomamkis and Craig, 1991; Craig
et af, 1992;).
Vanous preconcentration steps, such as mercury-gold amalgamation (Freimann and
Schmidt, 1982), contmuous flow, and on-line pretreatment have been used to improve
sensitivity and efficiency (Mateo et af, 1990; Munaf et af, 1990a; Welz et af, 1992).
On-line (Munaf et af, 1990b; Rapsomamkis and Craig, 1991; Craig et af, 1992) and off¬
line (Schroeder, 1989) separation techniques have been employed to facilitate mercury
speciation.
A variety of alternative techniques have been used to improve mercury detection
and speciation. Voltammetnc techniques have been enhanced by preconcentration
strategies (Daih and Huang, 1992) and electrode modification (Navratilova and Kula,
1992). Electrothermal atomization atomic absorption spectrophotometry has been used
after solvent extraction preconcentration (LeBihan and Cabon, 1990). Mercury speciation,
using chromatographic separations by high performance liquid chromatography
(HPLC)(Krull et af, 1986) and capillary gas chromatography (Kato et af, 1992), followed
by atomic emission detection has been described. Gas chromatographic (GC) techniques
for methylmercury determmation traditionally used electron capture detection (Horvat et
af. 1988) because of the sensitivity of the detector. Recently studied GC techniques
employ sample preconcentration (Lansens et af, 1990; Bulska et af, 1991) or headspace

14
injection (Lansens et af, 1989) coupled with microwave-mduced plasma (MIP) detection
(Lansens et aL, 1989; Lansens et aL, 1990; Bulska et aL, 1991) and inductively coupled
plasma-mass spectrometry (ICP-MS)(Shum et aL, 1992).
Mercury analyses m water and air historically have been flawed by contamination
that often exceeds ambient mercury concentrations (Fitzgerald and Watras, 1989).
Technical advances have incorporated new strategies for sampling (i.e. "clean sampling
technique")(Douglas, 1991), and detection limits have been lowered by implementation
of clean laboratory practices and improved analytical techniques (i.e. atomic fluorescence
spectrophotometry)(Bloom, 1989). Further, improved technology has enabled researchers
to quantify individual mercury species (Fitzgerald, 1986).
Atomic fluorescence spectrophotometry is a most promising and versatile
technique for mercury detection in environmental matrices, and is rapidly becommg
accepted as the standard technique for low-level mercury determinations (Bloom, 1989;
Tanaka et aL, 1992). Fluorescence technology is free from spectral interferences that
plague absorption technology (Churchwell et ah, 1987). Improved mercury detection
limits, furnished by cold vapor atomic fluorescence spectrophotometry (CVAFS), approach
0.6 pg Hg (0.003 ng L'1 for a 200 mL sample)(Bloom, 1989). Basic fluonmetnc
spectrophotometry (Mariscal et aL, 1992) has been used to improve the sensitivity of total
mercury determinations, while atomic fluorescence, following preconcentration and
chromatographic separation (Bloom, 1989; Tanaka et aL, 1992) permits mercury
speciation at very low analyte concentrations. Extensive speciation schemes that employ
"clean field and laboratory" procedures (Douglas, 1991), and improved separation and

15
analytical techniques (Wilken, 1992) broaden our ability to quantify mercury in
environmental matrices and to identify ecological transformations of mercury.
Improvements in atmospheric mercury determinations have incorporated
concentration steps, such as gold trap amalgamation (Barghigiani et ah, 1991), or selective
absorption tubes, to permit mercury speciation (Braman and Johnson, 1974; Schroeder and
Jackson, 1987). Neutron activation analysis after preconcentration (Albnnck and Mitchell,
1979) and LIDAR techniques have been described (Ferrara et ah, 1992).
The detection limits provided by traditional technology, such as cold vapor atomic
absorption spectrophotometry (CVAAS), have typically been sufficient for the
determination of mercury in soil, sediment and biological samples (Sullivan and Delfmo,
1982; Colma de Vargas and Romero, 1992). Systematic mercury contamination during
sampling and analysis does not usually influence the quantification of mercury m these
matnces. Microwave digestion (Navarro-Alarcon et ak, 1991) and gold amalgamation
preconcentration (Mudroch and Kokotich, 1987), respectively, speed sample preparation
and improve sensitivity of CVAAS technology. Separation techniques have been
employed to evaluate certain mercury species m environmental and biological samples.
For example, methylmercury can be determmed, after solvent extraction and subsequent
identification/quantification, usmg gas chromatography with electron capture detection
(GC-ECD)(Alvarez and Flight, 1984; Hight, 1987; Horvat et ak, 1990; Bulska et ak,
1991). While modifications of the GC-ECD method have employed improved extraction
procedures and analytical configurations (Lansens and Baeyens, 1990), organomercurials
have also been characterized with methods usmg HPLC (Hempel et ak, 1992; Stoeppler
et ak, 1992) and ICP-MS (Beauchemm et ak, 1988) technology.

16
Another approach to mercury speciation is to establish operational definitions that
categorize mercury species based on a common response to a senes of physicochemical
conditions (Schroeder, 1989). According to these speciation schemes, compound groups
are isolated by a vanety of sequential selective extraction procedures (Magos, 1971; Revis
et ah, 1990; Rapsomamkis and Andreae, 1991; Sakamoto et ah, 1992).
Much research must follow guidelines that are outlmed by state or federal agencies
(Winter et ah, 1977). As a consequence of regulated adherence to "standard methods,"
researchers are often limited by traditional analytical technology until the regulatory
agency accepts modified and contemporary technology. The expanding technological
advances for trace metal analyses, and the complexity associated with environmental
analytical chemistry (i.e. variable analyte concentration, speciation, and matrix
interference) necessitate that the researcher: 1) optimize a protocol for "self-evaluation"
m the laboratory, and 2) implement interlaboratory calibration studies that characterize the
utility of traditional and contemporary techniques when analyzing environmental matrices.
Comparative studies of parallel methods (Churchwell et ah, 1987; Horvat et ah, 1988;
Fnese et ah, 1990) and interlaboratory calibration studies (Thibaud and Cossa, 1989;
Cossa and Courau, 1990) allow researchers to: 1) compare methods and optimize routme
laboratory practices, 2) identify superior analytical activities (sampling, storage,
preparation, and analysis) and, 3) evaluate the quality of data provided by a particular
method or laboratory.
Cold vapor atomic absorption spectrophotometry (CVAAS) is suitable for total
mercury determinations m soil, sediment and biological tissue if measurements are

17
verified by the appropriate quality assurance/quality control measures (i.e. instrument
calibration against a standard reference material, and verification of instrument stability).
However, sample preconcentration or analytical modifications are essential for mercury
determmations of air and water samples. Low level mercury determmations in water and
air samples, using CVAAS technology, should be considered suspect until these data can
be compared with external determmations usmg alternative technology.
Global Mercury Cycle
The global cycle of mercury is mechanistically determined by its high vapor
pressure (2.4 x 10'3 mm Hg at 20°C)(Stewart and Bettany, 1982; Clarkson et af, 1984;
Schroeder et al, 1989). This unique physico-chemical attribute causes the global mercury
cycle to be distinctly different from that of other trace metals (Moore and Ramamoorthy,
1984) The global cycle of mercury, mvolvmg the solid, aqueous, and vapor phases, and
influenced by the stability of volatile mercury species, permits widespread and long-term
dispersion of this element. The global cycle of mercury is outlmed m Figure 2.1.
Mercury is released to the atmosphere from natural land and ocean degassing,
volcanic activity, and human-related activities. Particulate-phase mercury is deposited
readily from the atmosphere, while vapor-phase mercury enters the global atmospheric
cycle and is dispersed for long distances. Photo-oxidative processes and particulate-
scavenging mechanisms eventually convert the vapor-phase mercury mto a particulate
form that is deposited by dry deposition or is scavenged during precipitation events.

18
N
Atmosphere
c
D
L
M
Biosphere
H
â–º
â—„
Water
G
F X*
A
B
1
J
'
Sediment
E
-â–º
Geologic
Substrate
A. Assimilation
B. Decay
C. Decay
D. Assimilation
E. Metamorphism
F. Dissolution
G. Assimilation
H. Decay
I. Weathering
J. Mineralization
K. Volcanism
L. Evaporation
M. Condensation
N. Volcanism
Figure 2.1 Block Diagram of the Global Mercury Cycle

19
Mercury is transported from the land to aquatic environments by terrestrial
leachmg or by discharges associated with human-related activities. Sediment serves as
the primary smk for mercury as a result of the strong affinity of mercury for organic and
sulfidic substrates, although a fraction (<1%) of sediment mercury may be remobihzed
as labile mono- or dimethylmercury. Monomethylmercury biomagmfies m the food cham
and volatile dimethylmercury is released to the atmosphere. Natural terrestrial and
oceanic releases of mercury to the atmosphere (30-100 x 108 g Hg yr'1 and 20-100 x 108
g Hg yr'1, respectively) are roughly equivalent to anthropogenic atmospheric releases (20-
100 x 108 g Hg yr'1) (Fitzgerald, 1986, Kim and Fitzgerald, 1986).
Atmospheric mercury deposition to the terrestrial environment is estimated to be
(40-100) x 108 g Hg yr'1, while mercury deposition on the world ocean is estimated to be
(20-275) x 108 g Hg yr'1 (Fitzgerald, 1986). The broad estimates for the global mercury
cycle arise from the sparsity of reliable data for certain compartments of the environment
(Fitzgerald, 1986; Fitzgerald and Clarkson, 1991).
Global and Regional Interactions
In the atmosphere, particulate-phase mercury may be transported m a manner
similar to other metals (Nater and Gngal, 1992). For example, particulate emissions from
pomt sources such as volcanoes, fires, and industry, may establish regional mercury
gradients m the surrounding environment (Crockett and Kinmson, 1977; Lmdberg, 1980;
Fukuzaki et ah, 1986; Sengar et af, 1989; Barghigiani and Riston, 1991; Pfeiffer et af,
1991; Ferrara et aL, 1992). The transport of particulate-phase atmospheric mercury

20
depends on wind direction (Brosset, 1987). However, more than 95% of the total
atmospheric mercury inventory is in the gaseous elemental form, with an atmospheric
residence time of 0.7 to 2.0 years (Slemr and Langer, 1992).
Smce >95% of the total atmospheric mercury inventory is m the gaseous elemental
form, mercury accumulation in regional terrestrial and aquatic systems may be dictated
by global changes in the mercury cycle (Swain et ah, 1992). Conversely, local activities
may contribute readily to the global cycle.
Mercury deposition in Swedish soils has been linked to mercury emissions from
the United Kmgdom, Germany, and Poland (Hakanson et ah, 1990b) and 10-15% of the
mercury m fish from Swedish lakes has been attributed to mercury emissions from foreign
sources (Hakanson et ah, 1990a). Recent studies have attributed increases in sediment
mercury accumulation to increased global atmospheric mercury emissions (Meger, 1986;
Steinnes and Andersson, 1991; Swam et ah, 1992) corresponding to an estimated 2%
annual increase m the atmospheric mercury budget (Slemr and Langer, 1992). Natural
inputs of mercury to the global cycle include volcamsm (Barghigiani and Riston, 1991),
tectonic activity (Varekamp and Waibel, 1987), and ocean and land degassmg (Xiao et
ah, 1991).
Mercury in the Atmosphere
The total gaseous mercury (TGM) concentration (>99% Hg°) compnses greater
than 95% of the total atmospheric mercury component (Bloom and Watras, 1989). Total
gaseous mercury concentrations range between 1 and 7 ng m'3 for samples taken from the

21
Pacific Ocean, the Mediterranean Sea, Italy, and rural Wisconsin (Fitzgerald et al, 1984;
Ferrara et al, 1986; Fitzgerald et al, 1991)(Table 2.2). Given the sparsity of data, there
is no evidence for continental sources of TGM. Elemental mercury m the atmosphere has
a relatively long residence time (0.7 to 2.0 years)(Munthe and McElroy, 1992; Slemr and
Langer, 1992) and ambient concentrations are not significantly influenced by ram episodes
(Ferrara et al, 1986). Elemental mercury is eventually oxidized by a variety of chemical
oxidations (i.e ozonation)(Iverfeldt and Lmqvist, 1986) and photo-oxidative processes
(Munthe and McElroy, 1992) and the resulting ionic mercury species (Hg2+ and CH3Hg+)
are readily scavenged by rainfall (Iverfeldt and Lmqvist, 1982). Durmg ram episodes,
there is a washout event that delivers water-soluble mercury to the earth's surface (Ferrara
et al, 1986). Wet deposition is not a significant source of monomethylmercury to the
equatorial Pacific Ocean (Mason et al, 1992) and a Wisconsm seepage lake (Fitzgerald
et al, 1991), however. Bloom and Watras (1989) suggest that monomethylmercury
concentrations of 0.15 ng L'1 ([Hg]tottl= 2-5 ng L1) m precipitation m the northwestern
United States can account for most of the fish mercury budget in Washmgton state lakes.
Mercury in Water
Mercury can occur in natural waters in the elemental (Hg°), mercurous (Hg+I), or
mercuric (Hg+2) forms, depending on ambient pH, oxidation-reduction potential, and ionic
composition. The thermodynamic stability domains (EH-pH diagram) for predominant
compounds, under varying pH and redox conditions, are described m Figure 2.2.
Mercuric hydroxy- and chloro- complexes are favored under conditions of high ambient

22
Table 2.2. Atmosphenc mercury concentrations and stack emission concentrations.
Industrv/Source
[Hg], ng m'3
References
Pomt Sources Emissions
-Mercury smelter
6 x 10*
Albrmck and Mitchell, 1979
-Chlor-alkali plant
2 x 106
Albrmck and Mitchell,1979
-Coal-fired power plant
1 x 105
Germani and Zoller, 1988
-Coal-fired power plant
1 x 104
Albrmck and Mitchell. 1979
-Coal-fired power plant
2 x 103
Lmdberg, 1980
-Non-ferrous smelter
2 x 105
Albrmck and Mitchell. 1979
-Sewage sludge mcmerator
2 x 105
Albrmck and Mitchell. 1979
-Cement factory
7 x 10°
Fukuzaki et af, 1986
-Volcano (Mt. Etna, Sicily)
7 x 10°
Barghigiani and Riston. 1991
- Cinnabar deposit, (Mt. Amiata, Italy)
1 x 103
Ferrara et af, 1992
Ambient Atmosphenc Concentrations
- Italy and Mediterranean Sea
Ferrara et ah, 1986
Capraia Island (sea level)
6.8
Ferrara et aL, 1986
San Pelligrmetto, Italy (1000 m)
5.7
Ferrara et aL, 1986
Livorno, Italy (urban)
10.1
Ferrara et aL, 1986
R. Solvay, Italy (chlor-alkali)
22.5
Ferrara et aL, 1986
Mt. Amiata (cinnabar deposits)
16.4
Ferrara et af, 1986
- Mt. Amiata (10-20 m above ground)
2.5
Ferrara et af, 1992
- Equatonal Pacific Ocean
1.3
Fitzgerald et af, 1984
- North Central Pacific Ocean
1.8
Fitzgerald et af, 1991
- Little Rock Lake, WI
1.6
Fitzgerald et af, 1991
Precipitation
[Hg], pM
References
Italy and Mediterranean Sea
Ferrara et af, 1986
- Capraia Island (sea level)
96
Ferrara et af, 1986
- San Pelligrmetto, Italy (1000 m)
50
Ferrara et af, 1986
- Livorno, Italy (urban)
133
Ferrara et af, 1986
- R Solvay, Italy (chlor-alkali)
131
Ferrara et af, 1986
- Mt. Amiata (cinnabar deposits)
100
Ferrara et af, 1986
Northeast Pacific Ocean
45
Fitzgerald et af, 1991
Wisconsm, USA
52
Fitzgerald et af, 1991
Washington, USA
17
Bloom and Watras, 1989

23
oh
Figure 2.2: Thermodynamic stability diagram for mercury
(redrawn from Krabbenhoft and Babiarz, 1992)

24
pH and chloride concentrations. Mercury is readily complexed by high molecular weight
dissolved organic materials (humic and fulvic acids), typically associated with organic
sulfhvdryl moieties (Andren and Harnss, 1973; Mantorna et ah, 1978). Mercuric sulfides
are favored in reducmg environments (Lodemus et ah, 1987).
Mercury is delivered to aquatic systems from direct precipitation and terrestrial
runoff (Xiankun et ah, 1990). Typical mercury concentrations m fresh, estuarme, and
saline waters are presented (Table 2.3). In aquatic systems, mercury is readily adsorbed
to the surface of livmg and nonliving particulate material (Wilkmson et ah, 1989) due to
the strong adsorption capacity of organic particulates for mercury (Bilmski et ah, 1992).
Partition coefficients for mercury between suspended solids and water have been
calculated to be (1.34 - 1.88) x 105 (Moore and Ramamoorthy, 1984). These adsorption
and complexation processes increase the rate of mercury removal to the sediment via
particulate scavenging and sedimentation (Mantoura et ah, 1978; Wallace et ah, 1982;
Moore and Ramamoorthy, 1984). Durmg estuarme mixmg, increased salinity induces the
precipitation of mercury-humic complexes, and m saline environments, mercuric chloride
complexes may become a dominant mercury component (Morel et ah, 1975; Calmano et
ah, 1992).
The low concentrations of mercury in natural waters necessitate the use of clean
sampling techniques and contemporary analytical technology (Bloom, 1989; Douglas,
1991). Many traditional sampling and analytical techniques for mercury m natural waters
are confounded by errors due to contamination and/or analytical insensitivity (Fitzgerald
and Watras, 1989), hence, caution must prevail when evaluating historic data.

25
Table 2.3. Mercury concentrations determined for various natural water bodies.
Location
Type
[Hg], pM
Hg species
References
Swedish lakes
FW
0.8 - 2.0
methyl-Hg
1
FW
0.6- 1.3
methyl-Hg
FW
0.5 - 0.6
methyl-Hg
Little Rock Lake, WI
FW
0.7 - 2.9
reactive
2
Gironde Estuary, France
EST
21.8- 103.2
total
3
St. Lawrence Estuary,
EST
9.0- 15.0
dissolved
4
Canada
SW
2.4
dissolved
Pacific Ocean
SW
4.7 - 9.7
total
2
Alboran Sea, Spam
SW
0.2 - 0.7
reactive
5
Strait of Gibraltar, Spam
SW
0.2 - 0.6
reactive
5
East North Atlantic
SW
0.4- 10.0
total
6
English Channel, England
SW
1.0- 20.4
total
7
Bay of Biscay, France
SW
2.8 - 4.3
total
3
SW
1.4- 2.8
reactive
Nova Scotia
SW
2.2 (near-shore)
reactive
8
SW
2.3 (off-shore)
reactive
Adriatic Sea, Italy
SW
10.1 - 33.7
dissolved
9
SW
0.4 - 76.5
particulate
North West Atlantic
SW
3.3 - 4.7
total
10
North Central Pacific
SW
1.7- 2.5
total
10
North Atlantic
SW
4 (surface)
total
11
SW
10 (thermocline)
total
SW
<4 (deep)
total
North Pacific
SW
1.2 - 2.6
total
11
SW
1.0 - 1.8
total
Equatorial Pacific
SW
0.3 - 5.0
reactive
12
SW
0.1 - 1.0
DGM
SW
0.0 - 0.6
MMHg
SW
0.0 - 0.7
DMHg
Baltic Sea, Germany
SW
2.5
total
13
References: ‘Lee and Hultberg, 1990; 2Fitzgerald and Watras, 1989;
3Cossa and Noel, 1987; 4Cossa et al, 1988; ’Cossa and Martin, 1991;
6Cossa et al, 1988; 7Cossa and Fileman, 1991; “Dalziel, 1992;
9Ferrara and Maserti, 1992; 10Gill and Fitzgerald, 1987;
“Gill and Fitzgerald, 1988; “Mason and Fitzgerald, 1990;
“Schmidt, 1992
Abbreviations: SW (saline water), EST (estuarine water), FW (fresh water),
DGM (dissolved gaseous mercury), MMHg (monomethylmercury),
DMHg (dimethylmercury), total and reactive (total and reactive mercury)

26
Mercury in Sediment
Sediments are a primary sink for mercury in the environment (Tolonen et ah,
1988). Mercury concentrations found in contaminated and noncontaminated soil and
sediment are presented (Table 2.4). Mercury forms strong associations with organic
material m soil and sediment under aerobic conditions. In addition, under anaerobic
conditions, insoluble mercuric sulfides may form (Lmdberg and Harnss, 1974). Lindberg
and Harnss (1974) found that mercury in sediment porewater was associated with low
molecular weight (MW<500) dissolved organic matter in Florida Everglades sediment, and
with high molecular weight (MW> 100,000) dissolved organic matter m sediment from
Mobile Bay, Alabama.
Senaratne and Dissanayake (1989) hypothesized a mechanism by which dissolved
mercury, initially scavenged from the water column by organic particulate material, is
precipitated as mercuric sulfide after reducmg conditions are established in response to
the decomposition of sedimented organic matenal. Their estuarine studies further
suggested that mercury was readily adsorbed to the surface of detntal grams coated with
iron-manganese oxides. Controlled Experimental Ecosystem studies (Wallace et aL, 1982)
demonstrated the rapid removal of mercury from the water column, where more than 90%
of spiked mercury was associated with particulate, colloidal, and high molecular weight
organic materials in the sediment. Wmfrey and Rudd (1990) added 203Hg to an organic
sediment, demonstratmg that more than 99% of the radio-labelled mercury was retamed
by the substrate. Likewise, Krabbenhoft and Babiarz (1992) found that 92% to 96% of
deposited mercury was retamed by soils. The retention of mercury by organic soils is

27
Table 2.4. Mercury concentrations in variously impacted soil and sediment.
Location
[Hg], mg Kg'1
References
Little Rock Lake, WI sediment
0.10
1
Canadian peat
0.06
2
Okefenokee, GA peat
0.40
2
Lake Istokpoga, FL peat
2.45
2
Swedish soils
0.12 -
0.22
3
Sn Lanka tidal flat
4.40
4
Sn Lanka peat sediment
15.5
4
W. Everglades mangroves
0.22 -
1.86
5
Adriatic Sea
0.02 -
8.63
6
Municipal Wastewater Sludge
1.24
7
Sewage Treatment Plant (STP) study*
- Returned Activated Sludge (RAS)
12.3 -
30.0
8
- Non-Hg contaminated RAS
0.03
8
- Sediment (upstream of STP)
0.09
8
- Sediment (downstream of STP)
0.98
8
- Sewage-amended soil
14.6
8
- Sediment (chlor-alkali receivmg lagoon)
420.
8
'Fitzgerald and Watras, 1989; 2Roddy and Tomlinson, 1989; 3Steinnes and Andersson, 1991;
4Senaratne and Dissanayake, 1989; ’Lindberg and Harriss, 1974; 6Ferrara and Maserti, 1992;
Cappon, 1984; 801son et af, 1991

28
much stronger that for mineral soils due to the association of mercury with sulfhydryl
moieties m the organic materials (Barrow and Cox, 1992a, 1992b).
Barrow and Cox (1992a, 1992b) characterized the influence of ambient salinity on
mercury sorption. At low chloride concentrations, mercury sorption to organic material
was unchanged between pH 4 and 6, and decreased at pH values greater than 6 (Barrow
and Cox, 1992a). The maximum sorption of mercury to the mineral, geothite, occurred
at a pH less than 4 (Barrow and Cox, 1992b). At high chloride concentrations, mercury
sorption to organic material increased between pH 4 and 6, and decreased at pH greater
than 6 (Barrow and Cox, 1992a).
Complexation of mercury m humic-rich or saline waters may facilitate the
desorption of mercury from sediment (Lindberg and Harnss, 1974). Studies in the Krka
estuary of Russia suggested that mercury, associated with dissolved organic material in
the freshwater environment, was readily precipitated during estuarine mixmg, and that
mercury adsorption to mineral surfaces enhanced sedimentation at the freshwater/saltwater
interface (Bilmski et af, 1992). Mercury removal during estuarine mixmg likely
decreased mercury bioavailabilty m these regions (Calmano et ah, 1992).
Mercury from deep sediment may be released to overlymg water in particulate
form as a result of deep sediment crackmg that can occur during repeated drying and
flooding events (Lodenius et ah, 1987). Desorption of mercury from sediment can occur
under acidic conditions; however, typical sediment pH values are not sufficiently low to
elicit this response (Barrow and Cox, 1992a, 1992b). Strong bonds, between mercury and
sulfhydryl moieties of particulate organic matter, render mercury unavailable for uptake

29
(Langston, 1982, Duddndge and Wainwnght, 1991). Further, less than 1% of the
inventory of sediment mercury is present as methylmercury, thus largely unavailable for
assimilation (Henning et aL, 1989; Revis et ah, 1990; Wmfrey and Rudd, 1990).
Mercury in Biota
Various mercury levels have been identified m plant and animal tissues (Table
2.5). Mercury was shown to accumulate in the root tissue of Sp art in a altemiflora but was
not readily transported to the rhizomes and above-ground tissues (Breteler et aL, 1981).
Further, mercury accumulation rates did not increase for plants grown in soils that were
amended with sewage sludge fertilizers, but accumulation was inversely related to
sediment organic matter content (Breteler et aL, 1981). Elevated temperature and light
intensities mcreased mercury uptake in the aquatic macrophytes, Elodea densa (Maury -
Brachet et aL, 1990) and Ludwigia natans (Ribeyre, 1991). Fortmann et aL (1978)
suggested that plant mercury uptake can make deep sediment mercury available for
accumulation in the food chain or facilitate its return to the global cycle m the aqueous
or gas phases, or as detritus. Mercury abundance has been measured m plankton (Watras,
1993), invertebrates, fish (Barber and Whaling, 1984; Lange et aL, 1993), birds (Burger
and Gochfeld, 1991; Thompson et aL, 1992), raccoons (Roelke et aL, 1991), and top
carnivores (i.e. American alligator and Florida panther)(Roelke et aL, 1991; Heaton-Jones,
1992).
Historically, most studies evaluated mercury accumulation in animals because of
human health concerns. As a result, studies of mercury m consumables, such as sportfish,
dommate the literature (Schmitt and Brumbaugh, 1990). Growing ecological concerns

30
Table 2.5. Mercury concentrations in plant and animal tissues.
Sample [Hg], mg/Kg References
Marine Fish/Mammals
-Sardines 0.02
-Dolphms (Equat. Pacific Ocean) 1
-Pacific Blue Marlin 14.0
-Snapper 0.01
-Fish/Shellfish (Mmamata, Japan) 9.
Freshwater Fish
-Brown Trout 0.08
-Northern Pike 27.8
-Perch (Wisconsm lakes) 0.06
-Bass (Florida lakes) 0.03
-United States survey 0.01
-Walleye (Manitoba reservoir)8 0.2
-Walleye (Manitoba reservoir)b 0.5
Alligator (Florida Everglades)
-Farm-raised 0.10
-Native 1.50
Florida Panther (Florida Everglades) 100
Plankton
-Adriatic Sea 0.02
-Unknown 0.02
-Spam 0.50
Birds
-Starlings (United States survey) 0.01
Plants
-Lichens (Yugoslavia) 0.40
-Lichens (Finland) 0.06
-Fungi (Finland) 0.05
-Mosses (Finland) 0 04
-Ferns (Finland) 0.01
-Pines (Finland) 0.03
-Angiosperms (Finland) 0.01
-Angiosperms (USA)C 0.5
5
Beckert, 1978
Andre et af, 1990
1.66
Beckert, 1978
Chvojka et ah, 1990
24
Beckert, 1978
0.19
Beckert, 1978
Beckert, 1978
Cope et af, 1990
1.38
Lange et ah, 1993
0.37
Schmitt and Brumbaugh,
0.3
1990
Bodaly et af, 1984
1.0
Bodaly et af, 1984
(liver)
Heaton-Jones, 1992
Heaton-Jones, 1992
Roelke et af, 1991
0.14
Ferrara and Maseru, 1992
0.04
Mitra, 1986
16.80
Mitra, 1986
0.20
White et af, 1977
188.2
Lupsma et af, 1992
0.57
Nuorteva et af, 1986
1.40
Nuorteva et af, 1986
0.67
Nuorteva et af, 1986
0.06
Nuorteva et af, 1986
0.15
Nuorteva et af, 1986
0.22
Nuorteva et af, 1986
3.5
Mitra, 1986
“before flooding of reservoir,
bafter flooding of reservoir,
ctrees growing above cinnabar deposit

31
have resulted m the expansion of the scope of organismal mercury studies. Some studies
are geared toward understanding the role of food cham dynamics on biomagnification
(Watras, 1993). Florida panthers, for example, that fed primarily on raccoons exhibited
higher tissue mercury concentrations than those that fed on deer (Roelke et af, 1991).
Watras (1993) studied mercury in zooplankton from Little Rock Lake, Wisconsm, and
suggested that mcreased methylmercury production, resultmg from lake acidification,
resulted m mcreased mercury accumulation, and that bioconcentration factors (BCF) for
methylmercury species were related to the trophic level of the test organism.
Environmental Factors and Bioaccumulation
Mercury Transformations in Aquatic Systems
Methylation of mercury m the water and sediment of aquatic environments has
been shown to enhance the bioavailability of mercury to biota (Cope et ah, 1990; Watras.
1993). Numerous studies have generated a wealth of information regarding methylation
and demethylation of mercury m aquatic systems. However, the mechanistic complexity
of the aquatic mercury cycle confounds many attempts to characterize exclusive cause-
effect relationships among habitats (Miskimmm et ah, 1992).
The rate of mercury methylation is optimized under acidic and freshwater
conditions at an ambient temperature around 35°C (Bryan and Langston. 1992). Rates
of methylation decrease with mcreased pH and salinity, and are favored under moderately
anoxic conditions. Since methylmercury comprises less than 0.2% of the total mercury
concentration, it is unlikely that methylation plays a significant role m the post-

32
depositional migration of mercury, although methylmercury has been shown to play a
significant role m mercury bioavailability (Bryan and Langston, 1992)
Increased mercury release from anoxic sediment during lake stratification, with
subsequent precipitation by free sulfides, follows a pattern that is typical of the redox-
active metals (Bloom and Effler, 1990). However, the microbial (Berman et ah, 1990;
Choi and Bartha, 1993) and abiotic (Ebmghaus and Wilken, 1993) methvlation of mercury
in water and sediment adds another level of complexity to the mercury cycle.
Monomethylmercuric ion is readily assimilated into living tissue due to the propensity of
the mercuric ion to bind with sulfhydryl moieties of organic compounds (Hmtelmann et
ah, 1993). However, competitive mechanisms of assimilation, monomethylmercuric
sulfide precipitation, and sulfide-mediated disproportionation of monomethylmercuric ion
to volatile dimethylmercury, participate m a dynamic disequilibrium that influences the
compartmentalization of mercury between sediment, water, and biota (Bloom and Effler,
1990; Ferrara and Maserti, 1992). This disequilibrium is further driven by a variety of
environmental conditions including pH, alkalinity, organic content, and oxidation-
reduction potential (Steffan et ah, 1988; Winfrey and Rudd, 1990; Miskimmm et ah,
1992). For example, alkalme/anoxic conditions favor sediment dimethylmercury
production (Quevauviller et ah, 1992).
Abundant quantities of dissolved organic material dimmish methylation due to
mercuric ion chelation (Miskimmm et ah, 1992). Methylmercury production mcreases
with decreasmg pH from 7 to 5 at the sediment-water interface (Winfrey and Rudd, 1990;
Miskimmm et ah, 1992), but decreases with decreasmg pH m anoxic/subsurface sediment

33
(Steffan et af, 1988; Winfrey and Rudd. 1990). Volatilization of elemental mercury
decreases with decreasmg pH (Wmfrey and Rudd, 1990) and does not exceed 2% of the
ambient methylation activity.
Descriptive examples of the complexity of the mercury cycle m aquatic systems
include some antagonistic mechanisms imposed by physical and microbial conditions.
Miskimmin (1991) demonstrated that the solubility of monomethylmercury was directly
related to the dissolved organic carbon (DOC) content m natural waters. Although
sediment mercury methylation was not related to the DOC of overlymg water, the ambient
DOC facilitated the release of available monomethylmercury to overlying waters.
Interestingly, monomethylmercury(II)-DOC complexes were shown to dimmish mercury
bioavailability.
Sulfate reducmg bacteria have been identified as key participants in the
methylation of mercury (Berman et aL, 1990; Winfrey and Rudd, 1990; Oremland et ah,
1991; Choi and Bartha, 1993). While providing the mechanism to enhance
bioaccumulation via methylation, sulfate-reducmg bacteria produce sulfide, as a by¬
product of respiration. Ambient sulfides, produced by microbial sulfate reduction, may
effectively immobilize labile monomethylmercuric ion as monomethylmercuric sulfide
(Bloom and Effler, 1990). Interestingly, although more than 95% of methylation results
from cobalamin-mediated methylation by sulfate-reducers (Berman et af, 1990),
methylation rates were shown to mcrease under sulfate-limitmg conditions (Berman et af,
1990; Choi and Bartha, 1993). Under controlled condition, sulfate-reducmg bacteria
methylated less than 1% of available mercury under sulfate-reducmg conditions and

34
methylated approximately 40% of available mercury under sulfate-limitmg, fermentative
conditions.
Finally, demethylation processes provide another pathway for mercury speciation.
Monomethylmercury may be demethylated via an organomercunal lyase pathway m which
the covalent carbon-mercury bond is cleaved enzymatically, and the resultant mercuric ion
is reduced to elemental mercury with an enzyme-mediated mercuric reductase pathway
(Nakamura et af, 1990). Alternatively, methylmercury may undergo an oxidative
demethylation, in which monomethylmercury is used as an analog of a single-carbon
substrate for metabolism, with the concomitant production of carbon dioxide (Oremland
et af, 1991).
The competitive mechanisms of cobalamm-mediated methylation, and oxidative
and "organomercunal lyase"-mediated demethylation, establish domains under varying
environmental conditions. While aerobic demethylation m estuarine sediment appears to
proceed by the organomercunal lyase pathway, oxidative demethylation appears to be the
dominant pathway m anaerobic estuarine sediment and in anaerobic and aerobic
freshwater sediment (Oremland et af, 1991).
Ratios of methylation to demethylation (M/D) that are greater than one, m the
water column, demonstrate the dominance of methylation m the water column (Korthals
and Winfrey, 1987). Peak M/D ratios m surface sediment suggest that this
microenvironment may play a significant role m mercury bioavailability, while decreased
M/D ratios m deep sediment suggest that buried mercury, m the absence of bioturbation,
is rendered unavailable for biomagnification.

35
Bioaccumulation in Fish
Extensive studies have been carried out to identify the factors that influence
mercury accumulation m fish (Bodaly et af, 1984; Cope et af, 1990; Verta. 1990; Nilsson
and Hakanson, 1992). Mercury concentrations in fish are typically mversely related to
ambient pH, alkalinity, and primary production (Cope et af, 1990; Haines et af, 1992;
Lange et af, 1993), and are directly related to transparency (Lange et af, 1993). Fish
mercury concentrations have been shown to increase m newly flooded reservoirs (Bodaly
et af, 1984; Johnston et af, 1991). These increases result under new flooding conditions
when increased mercury methylation, a consequence of the microbial decomposition of
dead biomass, enhances the release of labile mercury from inundated soils.
Identification and Assessment of Mercury Contamination
Concerns over regional mercury contamination typically stem from discoveries of
local contamination. Publicity about mercury contammation m the Canadian provmce of
Saskatchewan, in the 1970s mcreased dramatically after the discovery of sediment
mercury contammation m the North and South Saskatchewan Rivers (Merkowsky et af,
1990). Subsequent studies were feverishly implemented to examine the cause and extent
of contammation. Evans (1986) studied mercury contammation in remote lakes of South
Central Ontario and quantified an average anthropogenic mercury loadmg (0.79 mg m'2)
to these lakes, with a final conclusion that the anthropogenic loadmg of mercury to these
remote lakes came from direct atmospheric deposition from outside the catchment area.
Once direct causal relationships of environmental mercury contammation can be identified

36
in studied systems, regional control measures may be implemented to ameliorate presumed
widespread contamination (Hamdy and Post, 1985).
Many questions remam regarding the abundance, transport, and cyclmg of
mercury, and the long term ecological effects of mercury contammation are uncertain
(Fitzgerald and Clarkson, 1991). However, sufficient information exists regarding the
contribution of anthropogenic activities to the global mercury budget (Clarkson et ah,
1984), the human health hazard, and major factors influencing mercury bioaccumulation,
to facilitate informed regulatory and remediation decisions.
Paleolimnological Studies
Paleolimnological studies suggest widespread recent and long-term mcreases of
mercury accumulation in aquatic systems, resultmg from changes m the global and
regional mercury cycle (De Lacerda et ah, 1991; Vincente-Beckett et ah, 1991; Swam et
ah, 1992; Wood et ah, 1992,). Decreases of mercury accumulation have resulted in
systems where mercury pomt sources have been eliminated (Cenci et ah, 1991). Regional
sediment mercury mcreases have been lmked to local mercury sources (Simóla and
Lodemus, 1982) and to global atmospheric mercury mcreases (Meger, 1986; Steinnes and
Andersson, 1991) .
Sediment cores from four lakes m Minnesota were dated radiochemically after y-
assay for unsupported 210Pb (Henning et ah, 1989). The resultmg mercury accumulation
rates suggested significant mcreases m mercury accumulation smce the turn of the
century. Pre-1850 mercury accumulation rates ranged from 4 to 15 pg m'2 yr'1, and

37
present mercury accumulation rates ranged from 10 to 100 pg m"2 yr'1. Mercury
concentrations in surface sediment were enriched 264% (80% to 450%, n=12) as
compared to deep sediment mercury concentrations.
Tolonen et ah (1988) dated sediment cores from the Baltic Sea near Oulu, Finland,
usmg the 210Pb dating technique. The resultmg age-depth relationship corresponded very
well with "varve" dates. Mercury accumulation rates from this location ranged from
approximately 50 pg m'2 yr-2 in 1920 to a peak accumulation rate (4200 pg nf2 yr'1) m
1970, with a declme to 700 pg m'2 yr'1 after 1980. The mid-century increases were
directly related to mercury discharges from a chlorine manufacturing facility.
Mercury accumulation rate profiles for dated sediment cores from seven lakes in
Wisconsm and Minnesota suggested that recent atmospheric mercury deposition rates were
3.4 times greater than those from pre-mdustnal times (3.7 to 12.5 pg m'2 yr'‘)(Swam et
al., 1992). The resultmg 2% average annual increase was compared to an estimated 1.5%
annual increase in atmospheric mercury concentrations over the North Atlantic Ocean
(Slemr and Langer, 1992) to suggest that global atmospheric mcreases were the primary
determinant of mercury accumulation m those aquatic systems (Swam et aL, 1992).
Norwegian sediment cores from ombrotrophic peat bogs showed increased mercury
concentrations from <50 ng g'1 in deep sediment (50 cm) to -190 ng g'1 in surface
sediment (Steinnes and Andersson, 1991). These mcreases corresponded to 188% (57%
to 363%, n=l 1) enrichment of mercury m surface sediment. Forest soil cores, receivmg
atmosphenc mercury inputs ongmatmg from a cement factory m Japan (Fukuzaki et af,
1986), suggested 183% (43% to 318%, n=5) enrichment of mercury m surface soils
compared to deep strata (40-50 cm).

38
Summary
Improved technology has enabled researchers to characterize the abundance,
speciation, and transport of mercury in the environment. As a result, current research,
has examined details of the global mercury cycle with greater interpretive resolution than
was previously possible.
Approximately half of the mercury inputs to the atmosphere are derived from
human-related activities. While particulate-phase emissions of mercury may establish
regional gradients of mercury in air, water, sediment, and biota, most mercury emissions
(90 to 99%) readily enter the global atmospheric mercury cycle as elemental mercury
vapor. Photo-oxidative and particulate scavenging mechanisms facilitate the conversion
of vapor-phase mercury to a particulate form that is subsequently deposited on the earth's
surface. Mercury forms strong associations with soil and sediment matrices, however,
biotic and abiotic processes facilitate the release of small quantities (<1%) of mercury to
overlymg water, primarily as aqueous monomethylmercury. Monomethylmercury readily
accumulates in biota and it’s concentration biomagnifies along the food web. At this time,
the complexity of the global mercury cycle, and the remaining technological barriers,
preclude a comprehensive assessment of ecological and human-health hazards imposed
by present and histone mercury contamination. However, the current understanding of
mercury behavior m the world ecosystem provides a strong foundation to compare and
contrast mercury contammation m environmental systems.

CHAPTER 3
MATERIALS AND METHODS
Site Selection
Sampling sites were selected to encompass a spectrum of conditions of
hydroperiod, soil type, and human impact (agriculture, urbanization), in an attempt to:
1) determme natural baseline mercury content and accumulation,
2) identify human-related changes in mercury content, accumulation,
and transport, and
3) characterize associations between mercury distribution and
selected physicochemical parameters.
Samples were retrieved from sites m seven major hydrologic regions descnbed as: Water
Conservation Areas 1, 2, and 3 (WCA), the Stormwater Treatment Areas (STA) within
the Everglades Agricultural Area (EAA), and the Everglades National Park (ENP)(Figure
3.1); the Okefenokee Swamp (OKE)(Figure 3.2); and Savannas State Reserve
(SAV)(Figure 3.3). The sampling regime was also chosen to optimize sampling of
transitional areas (i.e. agriculturally impacted vs. ummpacted). In regions with significant
water level variability, sediment cores were retneved from the wet areas rather than
nearby dry areas. Sediment cores were collected when possible, and soil grab samples
were collected m the few cases that a sediment core could not be obtained.
39

40
ar30- aroo' 80‘30' so’oo’
Figure 3.1. Sample locations in the Florida Everglades

41
82° 30’00” 82° 22’30 82°15’00" 82° 07'30"
0KE:56
0KE:57
0KE:58
Figure 3.2. Sample locations m the Okefenokee Swamp

42
2730’
2725’
2720’
2715’
80*30’ 80‘25’ 80'20’ 80*15’
Figure 3.3. Sample locations m the Savannas State Reserve

43
Field Sampling
Transport to sampling sites in the Water Conservation Areas, Everglades National
Park, and the Everglades Agricultural Area was provided by the South Florida Water
Management District usmg an airboat or a pontoon-equipped helicopter. The geographic
coordinates for sites accessed by helicopter were converted to latitude/longitude
coordinates from Global Positioning System (GPS) coordinates measured usmg on-board
equipment. Sample locations m the Savannas State Reserve (SAV) and the Okefenokee
Swamp (OKE) were accessed by foot, or by canoe. Sample coordmates for these
locations were determmed usmg quadrangle maps (latitude/longitude).
Temperature, conductivity, and dissolved oxygen of surface waters were measured
in situ usmg YSI (Yellow Springs Instruments) portable field meters and pH was
measured usmg a Fisher Scientific Accumet portable pH meter. Sediment cores were
obtained usmg thick-walled polyvmyl chloride (PVC) tubmg (7.5 cm diameter, 80 to 100
cm length). Core barrels were inserted slowly mto the sediment matrix to minimize
compaction. Once inserted, the top of the core barrel was capped with a large rubber
stopper. The core barrel was maneuvered from side to side and then pulled from the
substrate. The bottom of the core was then sealed with a large rubber stopper and the top
stopper was removed to fill the top of the core barrel with water to reduce any movement
of the sediment. The top rubber stopper was then replaced and both stoppers were taped
securely with duct tape. Cores were transported upright to the base camp for extrusion.
Sediment compaction averaged 27% (17-36%, n = 9).
For extrusion of the sediment, core tubes were attached to a vertical galvanized
pipe. A piston was inserted mto the bottom of the core barrel. The core barrel was

44
lowered while the piston was held stationary and two centimeter sections of sediment
were removed, sequentially, from the top of the core barrel. The core extrusion was
continued until the entire core was sectioned from the surface to deeper strata. Core
sections were transferred to previously labelled Whirlpak bags. Sample bag labels
included the sample identification number, date of sampling, and the initials of personnel
involved with core extrusion. All sediment samples were stored m the dark at 4°C in an
insulated chest during field operations and transported to the laboratory. Samples were
then placed in a freezer until sample analysis was initiated.
Total Mercury
Total mercury was determmed usmg the digestion procedure described in EPA
method 7471 for the determmation of mercury in soil and sediment followed by cold
vapor atomic absorption spectrophotometry (U.S.E.P.A., 1986). Sediment samples were
mixed m the Whirlpak sample bags, usmg an acid rmsed teflon-coated spatula, and two
grams of wet sample were transferred and weighed (to 0.0001 g) into a 10 mL plastic
beaker cup on a Mettler AE-160 analytical balance. The sample was transferred
quantitatively to an acid-rmsed 300 mL BOD bottle with a 10 mL deionized water rinse.
The digestion mvolved addition of 2.5 mL of concentrated nitric acid and 5 mL of
concentrated sulfuric acid. The sample was heated at 95°C for two mmutes. then 15 mL
of potassium permanganate (50 g L'1), and 8 mL of ammonium peroxydisuifate (50 g L*1)
were added to the digestion mixture. The sample was then heated at 95°C for one hour.
An additional 15 mL of potassium permanganate solution was added to the digestion

45
mixture if the permanganate color disappeared within fifteen mmutes of the initial
addition. Upon completion of digestion, samples were cooled and decolorized by the
addition of 6 mL of hydroxylamme hydrochloride solution (120 g hydroxylamine sulfate,
and 120 g sodium chloride per liter of deionized water).
Each digested sediment sample was transferred to a plastic reaction vessel fitted
for a Perkm Elmer MHS-10 cold vapor unit. Stannous chloride solution (80 g L'1) was
added contmuously (10 mL per minute) to the digestate m the reaction vessel. The
sample was contmuously purged with high purity nitrogen gas. Elemental mercury was
evolved from the digestate and swept with the nitrogen purge-gas into an open ended
quartz tube (1 cm diameter) with a 16 cm cell path length. The mercury was quantified
by cold vapor atomic absorption spectrophotometry usmg a Perkin Elmer model 5000
Atomic Absorption Spectrophotometer (A,=253.6 nm, SBW=0.7 nm) with a mercury
hollow cathode lamp (1=6 mA). Light absorption was measured as peak height. The
standard calibration curve workmg range (0 to 50 ng Hg) gave an absorbance range from
0.003 to 0.035 absorbance units. The detection limit for mercury analysis was 10 ng g'1.
Percent Solids/Bulk Density
Percent solids were determined by weighing known volumes of wet sediment m
aluminum weighing dishes. Wet sediment was transferred mto a 25 cm3 glass syrmge that
was modified to function as a piston chamber. The empty dish was weighed, a known
volume of wet sediment was transferred to the dish and weighed. The sample was dried
m an oven for 24 hours at 104°C, removed, and placed m a desiccator for approximately

46
1 hour. The dried sample was then re-weighed. The wet and dry sample weights were
corrected for the weight of the empty dish. Percent solids were then calculated as the
percent of dry mass to total wet mass. Bulk density was determined from the same
aliquot of wet sediment. The dry bulk density of the sample was calculated as the dry
sediment mass per 10 cm3 sample volume (g cm'3).
Radionuclide Analysis
To calculate age/depth relationships m sediment cores, the activity of unsupported
210Pb was estimated by determining total and supported 210Pb activity. Supported 2l(,Pb
results from, and is maintained by, radioactive decay of 226Ra (half-life 1622 years) in the
sediments. Unsupported 210Pb is formed by decay of 226Ra to 222Rn (half-life 3.8 days),
which escapes to the atmosphere, decays to 210Pb, and is deposited to sediment via
precipitation. Subtractmg supported 210Pb from the total measured activity of 210Pb m
sediment samples yields the unsupported 210Pb activity, that will decrease with depth in
the sediments because of radioactive decay. The age of a sediment layer may then be
calculated from its activity of unsupported 2l0Pb. Because the half-life of 210Pb is only
22.3 years, this datmg technique is restricted to about a 150-year time span. Activity of
137Cs serves as an independent age marker because it first appeared m the atmosphere
during nuclear bomb testmg around 1960.
Activities of 210Pb and l37Cs were measured by direct y-assay usmg two mtnnsic-
germanium well-detectors (Princeton Gamma Tech). This type of detector counts over
a large range of y-energies and can be used for simultaneous measurement of supported

47
and unsupported 210Pb (Gággeler et ak, 1976), as well as l37Cs which may be used as an
additional age-marker (Ritchie et ak, 1973). Lead shielding (10.1 cm thick) was used to
reduce natural background radiation at the germanium detector. Samples for radionuclide
analysis were dried at 95°C for 24 hours, pulverized by mortar and pestle, weighed, and
placed m small low-density polypropylene tubes (capacity 4 mL). The volume of the
samples and standard were matched to ensure the same counting efficiencies for both.
Core sections were combmed (up to 2 cm) to obtain an adequate sample volume. Sample
tubes were sealed with plastic cement and left for a minimum of 14 days to equilibrate
radon (222Rn) with radium (226Ra).
Counting times varied from 7 to 26 hours depending on sample weight; small
samples needed longer counting times to minimize uncertainty. For each region of
interest, counts were corrected for Compton scattering by subtractmg the below-the-peak
area from the total counts. This area was determmed by a linear fit through three channel
contents (e g. counts) on either side of the region of mterest.
Blanks were counted for every two samples to determine background from ambient
radiation. Standards (Department of Energy, New Brunswick Laboratories: U-Th
standards) were run with the same frequency to track efficiency (counts y'1) and to
calculate a 226Ra conversion factor (pCi counts'1 s"1). Sample spectra were analyzed for
activity in the 46.5 keV (210Pb) and 662 keV (137Cs) peaks. Activities at 295 keV (214Pb),
352 keV (214Pb), and 609 keV (214Bi) representing uranium series peaks were used to
compute supported 210Pb abundance.

48
Calculation of 210Pb dates followed the constant rate of supply (CRS) model
(Goldberg, 1963) which is able to quantify changing sediment accumulation rates. This
model appears applicable to Florida aquatic systems, particularly because :i0Pb residuals
match both the known atmospheric flux of this isotope as well as the residuals of nearby
cores (Bmford and Brenner, 1986; Gottgens, 1992). These residuals are defmed as the
total inventory of unsupported 2I0Pb (pCi cm'2) m the core from the surface to the depth
at which its activity has decayed to background levels. Such a constant rate of 2l0Pb
fallout is likely, due to the high efficiency at which 2I0Pb is scavenged from the
atmosphere and from the water column by wet precipitation or particulate matter
(Turekian et ak, 1977; Robbms, 1978). This provides evidence favormg the assumption
of the CRS dating model that an increase in the rate of delivery of bulk sediments will
not supply more 210Pb. Finally, a constant rate of 210Pb fallout will result in different
unsupported 210Pb activities at the sediment-water interface m core locations with differing
rates of net sediment accumulation. This has been confirmed by paleolimnological
investigations in aquatic systems throughout Florida (Bmford and Brenner, 1986).
Uncertainty analyses were based on both the random variation of counting errors
associated with radioactive decay and the nature of the CRS model. Errors controlled by
external forces such as maccuracies of stratigraphic sampling and determination of bulk
density were not considered.
Radiation emitted m nuclear decay is subject to statistical fluctuation. This
unavoidable source of uncertainty is often a predominant source of imprecision (Knoll,
1979). Because the recorded counts m nuclear counting experiments follow a Poisson

49
distribution, the predicted standard deviations were estimated as the square root of the
mean number of counts. The amount of 210Pb (total, supported, and unsupported) and
l37Cs was expressed as activity (pCi g'1) ± one standard deviation (i.e. 68.3% confidence
limits), which is standard practice in expressing uncertainty in nuclear measurements
(Wang et ah, 1975; Binford, 1990). Counting errors m the calculation of net ísotope-
activities were propagated usmg first-order analysis.
Monte Carlo simulation (Palisade Corp., 1990) was used to estimate error
associated with the calculation of age and sedimentation rate following the CRS model.
The probability density function for simulated 210Pb activities was approximated by a
normal distribution with the mean equal to the measured activity and a range equal to the
counting error.
Carbon
Total Carbon
Total carbon was analyzed usmg a Coulometer (Coulometncs, Inc., Model 5011)
combmed with a Total Carbon Combustion Apparatus (Coulometncs, Inc., Model 5020).
Total carbon measurements were made by weighing approximately 5 mg of air dned
sediment into a platmum boat. The platmum boat, contammg the dned sample, was
placed in the entry port of a preheated (950°C) furnace. Contaminant C02 was swept
through the furnace to the attached coulometnc cell. The coulometer solution was titrated
coulometncally to eliminate contaminant interference. The platmum boat, contammg the
sample, was then moved from the entry port into the furnace. Carbon dioxide, evolved

50
from the sample, was swept mto the coulometnc cell. The resultant pH change mduced
a color change m the coulometnc solution. The solution was then titrated coulometrically
to the initial pH and color. The analysis quantified in units of micrograms carbon and
percent of total carbon was calculated.
Inorganic and Organic Carbon
Inorganic carbon was analyzed usmg the coulometnc procedure, described above,
coupled with a Carbonate Carbon Apparatus (Coulometrics, Inc., Model 5030). Dry
sediment (10 to 20 mg) was transferred to a porcelain boat, placed m a glass tube, and
attached to the Carbonate Carbon Apparatus. The glass tube was placed on a heatmg
element and 3 mL of perchloric acid (2 N) was mtroduced to the sample. Carbon
dioxide, evolved from the sample, was swept mto the coulometnc solution and titrated
(Huffman, 1977; Lee and Macalady, 1989). Organic carbon was determined as the
difference between the total and inorganic carbon content.
Additional Trace Metals
Analyses for cadmium (Cd), chromium (Cr), copper (Cu), iron (Fe), nickel (Ni),
lead (Pb), and zinc (Zn), were performed on 0.5 to 1 gram dned sediment aliquots usmg
the digestion procedure described m EPA Method 3050 (U.S.E.P.A., 1986). Ten mL of
1:1 nitric acid (HN03) were added to the sediment m a beaker and covered with a watch
glass. The mixture was heated to 95°C and refluxed for 10 to 15 minutes without boilmg.
The sample was cooled and 5 mL of concentrated HN03 was added. The watch glass

51
was replaced and the solution was allowed to reflux for 30 minutes. The last step was
repeated to ensure complete oxidation. Covered with the watch glass, the solution was
then concentrated by evaporation to 5 mL without boilmg. Two mL of deionized (DI)
water and 3 mL of 30% hydrogen peroxide (H202) were added to the cooled solution.
The beaker was covered with the watch glass and was warmed on the hot plate to initiate
the peroxide reaction. Hydrogen peroxide was added in 1 mL aliquots, with warmmg,
until effervescence became minimal or until the general sample appearance was
unchanged. Not more than 10 mL of 30% H202 were added to minimize acid dilution
and digestate volume. Next, 5 mL of concentrated hydrochloric acid (HC1) and 10 mL
of DI water were added to the solution, covered and returned to the hot plate to reflux for
an additional 15 mmutes without boilmg. After coolmg, the solution was filtered through
Whatman No. 41 filter paper to remove particulates. The filtrate was diluted to 100 mL
with DI water. The acid concentration was 5.0% (v/v) HC1 and 5.0% (v/v) HN03 for the
diluted solution.
Metals were quantified, by flame atomic absorption spectrophotometry (FAAS),
using a Perkin Elmer model 5000 Atomic Absorption Spectrophotometer, with appropriate
hollow cathode lamps and an air/acetylene flame. The following instrument settings
(Table 3.1) and detection limits (Table 3.2) were determined.

52
Table 3.1. Instrument settings for metal analyses usmg a Perkm
Elmer model 5000 Atomic Absorption Spectrophotometer
Element
Wavelength
nm)
Bandwidth
(nm)
Cr
357.9
0.7
Pb
283.3
0.7
Ni
232.0
0.2
Cd
228.8
0.7
Zn
213.9
0.7
Cu
324.7
0.7
Fe
248.3
0.2
Table 3.2. Detection limits for metals determination usmg a Perkm
Elmer Model 5000 Atomic Absorption Spectrophotometer (Flame
Atomizer)
Analyte
Detection Limit
(mg/Kg dry weight)
Cd
2
Cr
5
Cu
2
Fe
8
Ni
5
Pb
16
Zn
1

CHAPTER 4
RESULTS AND DISCUSSION
Water Quality
Sample site coordinates (latitude/longitude) and associated water quality parameters
(depth, temperature, conductivity, dissolved oxygen, and pH) are presented in Table 4.1
These data demonstrate the variability of water quality and quantity throughout the
Everglades region. Water depth at Everglades sites (ENP, WCA1, WCA2, WCA3, and
STA) ranged from 0 to 0.6 meters. Water depth at the Savannas sites (SAV) ranged from
0 1 to 1.4 meters Okefenokee sites (OKE) were covered by a floating Sphagnum spp.
mat (approximately 0.5 m thick) with approximately 0.5 meter of underlying water.
Conductivity of overlymg water ranged from 49 to 37000 pmhos cm'1 for the
Everglades. Average conductivities for WCA1, WCA2 and WCA3 were 257, 1257, and
625 pmhos cm'1, respectively, while those for the ENP were regionally variable, with a
measured range from 465 to 37000 pmhos cm'1. Water at the periphery of WCA1 is
supplied to some degree by agricultural runoff and exhibits higher conductivities (230 to
850 pmhos cm'1), while most of the water in the center of WCA1 is derived from
precipitation (49 to 98 pmhos cm'1). The conductivities of water at OKE and SAV
sample sites also mdicate the predominance of precipitation to the regional hydrology
(OKE, 42 to 121 pmhos cm'1; SAV, 54 to 74 pmhos cm'1, respectively).
53

54
Table 41. Water quality data associated with wetland sediment sample sites
Sample ID#
Latitude
Longitude
Depth
Temp.
PO]
Cond.
pH
(m)
(deg. C)
(mg/L)
(umhos/cm)
ENP :01
254303
804311
0.10
24.5
6.4
465
8.0
ENP :02
254121
803809
0.10
26.0
2.8
780
7.6
ENP: 03
253101
803802
0.00
N/A
N/A
N/A
N/A
ENP: 04
253647
804129
0.10
26.0
5.9
500
7.9
ENP :05A
252004
804450
0.00
N/A
N/A
N/A
N/A
ENP :05B
252004
804450
0.00
N/A
N/A
N/A
N/A
ENP :05C
252004
804450
0.00
N/A
N/A
N/A
N/A
ENP :05D
252004
804450
0.00
N/A
N/A
N/A
N/A
ENP :05E
252004
804450
0.00
N/A
N/A
N/A
N/A
ENP :05F
252004
804450
0.00
N/A
N/A
N/A
N/A
ENP :06
251457
803608
trace
30.0
1.0
8200
7.7
ENP: 07
251705
803805
0.05
27.0
2.0
500
7.6
ENP :08A
252754
805114
0.00
N/A
N/A
N/A
N/A
ENP :08B
252754
805114
0.00
N/A
N/A
N/A
N/A
ENP :09
253625
811014
0.03
21.8
1.0
37000
7.2
ENP: 10
253201
810011
trace
25.0
4.3
23000
7.1
ENP: 11
253119
804741
0.10
26.0
3.5
1000
7.4
ENP: 12
253632
805632
0.13
26.0
5.0
625
7.4
WCA3:13
255024
804944
0.45
27.0
6.8
325
7.8
WCA3:14
254959
804156
0.45
26.0
5.6
405
7.4
WCA3:15
254953
803305
0.15
28.5
8.5
700
7.8
WCA3:16
255707
802905
0.15
25.5
2.4
750
7.5
WCA3:17
255702
804151
0.45
26.0
4.2
480
7.4
WCA3:18
260400
803805
0.30
21.5
1.5
460
7.4
WCA3:19
260401
804804
0.30
21.5
3.8
500
7.2
WCA3:20
261802
804754
0.00
N/A
N/A
N/A
N/A
WCA3:21
261014
804457
0.05
ND
ND
ND
ND
WCA3:22
260914
804201
0.30
23.0
2.4
650
7.5
WCA3:23
261756
803652
0.10
25.0
7.9
900
6.0
WCA3:24
261002
803302
0.15
25.0
3.0
800
7.2
WCA2:25
261041
802156
0.15
25.0
2.0
900
7.1
WCA2:26A
261800
802056
0.15
25.0
6.3
1350
7.3
WCA2:26B
261800
802056
0.15
25.0
6.3
1350
7.3
WCA2:27
262555
802652
0.10
15.5
7.2
1200
7.5
WCA2:28
261901
802658
0.15
17.0
2.7
1325
7.4
WCA2:29
262149
802058
0.05
16.0
1.9
1350
7.8
WCA2:30
262034
802030
0.30
17.0
1.8
1250
7.3
WCA2:31
261954
802105
0.15
17.5
1.5
1425
7.3
WCA3:32
260147
802855
0.45
19.5
2.3
800
7.3
WCA3:33
255923
803053
0.60
19.8
2.7
700
7.3
N/A corresponds to sites with no overlying water
ND corresponds to data not determined

55
Table 4.1. (cont'd)
Sample ID#
Latitude
Longitude
Depth
Temp.
[DO]
Cond.
pH
(m)
(deg. C)
(mg/'L)
(umhos/cm)
WCA3:34
255739
803219
0.45
20.5
2.5
650
7.3
WCA1:35
264005
802141
0.30
26.5
0.5
850
6.7
WCA1:36
263449
802047
0.20
29.5
2.7
90
7.0
WCAL37
262924
801939
0.55
30.5
2.1
82
6.6
WCA1:38
262806
802441
0.55
30.0
1.8
442
6.7
WCA1:39
263200
802447
0.10
30.0
3.0
230
7.1
WCA1:40
262258
801657
0.25
31.0
1.6
49
7.5
WCA1:41
262719
801452
0.30
30.5
1.3
98
7.4
WCA1:42
263304
801543
0.35
34.2
1.8
218
7.6
ST A :43
263919
802510
0.30
35.0
1.2
530
7.5
ST A :44
263736
802526
0.00
N/A
N/A
N/A
N/A
ST A :45
263854
802440
0.10
32.0
2.3
500
8.5
ST A :46
263842
802537
0.30
34.0
0.5
560
7.5
STA :47
263927
802436
0.00
N/A
N/A
N/A
N/A
SAV :48
271630
801500
1.4
20.2
6.2
114
ND
SAV :49
271645
801530
1.1
22.4
7.6
121
ND
SAV: 50
272115
801830
0.00
N/A
N/A
N/A
N/A
SAV :53
272000
801750
1.0
18.0
ND
42
ND
SAV :54
272015
801730
1.0
18.0
ND
72
ND
SAV :55
271945
801700
0.1
17.8
ND
78
ND
OKE : 56
304235
821000
0.5
17.0
2.4
74
ND
OKE :57
304235
821000
0.5
17.2
5.3
68
ND
OKE :58
304235
821000
0.5
19.0
7.2
54
ND
N/A corresponds to sites with no overlying water
ND corresponds to data not determined

56
High conductivity water m Everglades National Park (ENP) is related to estuarine
mixing while high conductivity water at Water Conservation Area (WCA) sampling sites
is an indicator of hydrologic inputs from the Everglades Agricultural Area (EAA). Low
conductivity of surface water in the center of WCA1 has been used to demonstrate that
the primary hydrologic source is from precipitation (Richardson et ah, 1990; SFWMD,
1992).
Sediment Geochronology
Results of paleolimnological analyses may be presented m units of concentration
or as rates of accumulation. Concentration, expressed as a relative measure of sediment
composition (e.g. mg g1), is the conventional way of expressing sediment stratigraphy
(Shapiro et ah, 1971; Pennington, 1973; Griffiths and Edmondson, 1975). Such data,
however, are vulnerable to variations in sedimentation of other components in the profile.
These variations may result in dilution of the target analyte. This problem can be
eliminated by using ratios of components in the sediment matrix, or by calculating
accumulation rates. The latter are normalized to time thus avoiding the problem of co¬
variance among sedimentary components.
Compilations of all data for sediment cores retrieved from the Everglades,
Okefenokee Swamp, and Savannas State Reserve appear m the Appendix. Blank cells m
the tables of the Appendix occurred so that all data could be presented m a consistent
tabular format. The tables include total mercury, solids, bulk density, total and organic
carbon, cadmium, copper, chromium, iron, lead, nickel, and zinc. These data are
summarized m Table 4.2. The Appendix also includes aspects of sediment geochronology

Table 4.2. Average concentrations of parameters measured m soils from the Water Conservation Areas (WCA1, WCA2, and WCA3),
Everglades National Park, Savannas State Reserve, and the Okefenokee Swamp
[Hg]
(ng g'1)
Solids
(8dry §wet )
Carbon
Total Organic
(%) (%)
[Cd]
[Cr]
[Cu]
[Fe]
- mg Kg'
[Ni]
1
[Pb]
[Zn]
WCA1
Recent
243
0.010-0.159
43.5
43.9
3
6
20
2257
6
53
36
Historic
81
0.044-0.137
47.3
48.7
2
5
10
1546
5
27
11
WCA2
Recent
155
0.013-0 167
41.6
41.5
3
6
12
2600
5
42
23
Historic
47
0.049-0.190
45.7
45.8
2
5
12
2813
5
25
13
WCA3
Recent
99
0.056-0.235
40.1
42.4
2
5
11
8385
5
58
43
Historic
55
0.080-0.648
43.5
43.9
2
5
6
6970
5
22
9
ENP
Recent
67
0.075-0.283
28.6
14.4
4
14
13
9060
13
77
22
Historic
44
0.103-0.542
24.9
15.6
4
24
12
10353
13
55
6
SAV
Recent
98
0.130-0.373
Historic
42
0.132-0.670
OKE
Recent
98
0.047-0.091
Historic
73
0 058-0.1 11
* "Recent" soil is
classified as
the summation of all "
post-1985"
strata (i.e.
0-4 cm).
* "Historic" soil is classified as the summation of all
"ca. 1900"
strata (i.e.
11-17 cm).

58
based on 210Pb dating (i.e. sediment accumulation rate, mercury accumulation rate, and
age/depth relationships). Total mercury, percent solids, bulk density, and water quality
were measured for all sample sites, while sediment geochronology, total and organic
carbon, and additional metals were determined for selected samples.
Sediment Datmg Acceptance Criteria
Radiochemical techniques are routinely used to date lake sediment profiles. Few
attempts have been made, however, to apply these methods in wetland sediment.
Diagenesis in wetland deposits is poorly understood and the correlation between depth and
time-of-deposit may be affected by compaction, decomposition, and vertical migration of
the element for which sedimentation rates are computed. Compaction may be accounted
for by calculating material deposition in units of mass (grams) rather than depth (cm) over
time. Decomposition of organic matter may increase the concentration of the analyte of
concern (e g. 210Pb, mercury, and others). Because of the nature of the CRS model,
however, core sections with such concentrated 210Pb (C) will have proportionally lower
deposition rates for bulk sediment (r) and, thus, for the analyte of concern (its
concentration multiplied by the bulk sedimentation rate). This follows from the CRS-
calculation for sedimentation rate according to
r
k A
C
(1)
where
r = bulk sediment accumulation rate (g cm'^yr1)
A = the residual 210Pb beneath the sediment horizon of interest (pCi cm'2)
k = 210Pb radioactive decay constant (yr'1), and
C = unsupported 210Pb activity in the sediment honzon of interest (pCi g'1).

59
The potential for temporal variability in the depositional environment in a wetland
may limit the resolution of a core's age/depth profile. Bummg of dry, organic wetland
soil, for example, may cause a loss of 210Pb from the profile to the atmosphere (fly-ash).
This reduces the cumulative residual 2l0Pb, i.e. the amount of this isotope (pCi cm'2) in
the core from the surface to the depth at which its concentration has decayed to
background level. Such reduction makes age/depth determinations less accurate.
Confidence in the dated profiles is enhanced when cumulative residual 210Pb corresponds
among cores from the same area despite differences in sediment accumulation rates.
The cores analyzed from this 5600 km2 Everglades area showed an average 2I0Pb
residual of 15.5 pCi cm'2 (sd = 3.5, n = 20). The range of residuals corresponded to 210Pb
fallout rates between 0.33 and 0.67 pCi cm'2 y'1, which was well within the normal range
of 210Pb fallout of 0.2-0.9 pCi cm'2 y'1 (Appleby and Oldfield, 1983). Seven cores with
fallout rates outside this range were excluded from the analysis (Figure 4.1). These
profiles may have been disturbed over time by removal or addition of material (producmg
a lower or a higher cumulative residual 2l0Pb, respectively).
Additional support for age/depth relationships may come from matchmg peak-137Cs
activity m the profile with a 210Pb-determined age of 1963 (Krishnaswami and Lai, 1978).
These peaks (or the onset of 137Cs activity m the absence of a distmct peak) occurred in
the profiles (n = 18) at an average "l0Pb-determined age of 1962, although the range of
age-values was considerable (1942-1978) (Figure 4.2). This may suggest some post-
depositional mobility of 137Cs (up or down m the core). An additional two cores
(WCAL36; WCA3:17) were excluded from consideration because their assigned dates to
peak l37Cs activities fell outside this range (Figure 4.2).

60
"2®Pb Fallout Rate (pCi cm ^yr 3)
oootN^rcocoo co
— — — — — CNI CN
Cumulative Residual Opb (pCi/cm^)
Figure 4.1 Cumulative residual unsupported 2I0Pb (pCi cm'2) for all cores
analyzed radiochemically. Cores with fallout rates outside the range 0.33-0 67 pCi
cm ‘ vr'1 were not mcluded m the computation of material accumulation rates.
Fallout of 210Pb is the product of the 2l0Pb residual and the radioactive decay
constant for 210Pb. Core identifications are placed within the bars
O
0
Q_
(/)
e>
rO
O
0
cn
o
TD
0
C
£
0
*0
“O
-O
CL
O
CN
2000
1980
1960
1940
1920
1900
1880
WCA 1 WCA 2 WCA 3 ENP SAV
Figure 4 2. '10Pb detenmned age of the core section with peak activity of l37Cs
(•) for dated sediment cores from Water Conservation Areas 1.2. and 3.
Evergiades National Park; and Savannas State Reserve. Open data pomts (O)
represent profiles m which the onset of l37Cs activity was used as a marker
horizon m the absence of a distinct 137Cs peak.

61
Sediment Mercury Geochronology
The geochronology of cores that satisfied the above described sediment datmg
acceptance entena are presented in Figures 4.3 through 4.20. Recent (post-1985) and
histone (approximately 1900) average sediment accumulation rates for each sample region
are given m Table 4.3. The mercury accumulation rate is calculated as the product of the
sediment accumulation rate and the total mercury concentration at each depth interval of
the sediment profile. Turn of the 20th century (ca. 1900) and recent (post-1985) mercury
accumulation rates for dated cores are averaged by sample region (Table 4.4).
Table 4.3. Recent and histone average sediment accumulation rates m cores retneved
from Water Conservation Areas 1,2, and 3, Everglades National Park, and Savannas State
Reserve. Numbers m parentheses indicate the range of values found.
Sample
Number of
Average Sediment Accumulation Rate (g cm'2 y'1)
Region
Cores
1900
Post-1985
WCA1
5
0.018 (0.009-0.030)
0.047 (0.016-0.099)
WCA2
3
0.021 (0.011-0.030)
0.042 (0.031-0.064)
WCA3
3
0.015 (0.009-0.023)
0.069 (0.029-0.143)
ENP
5
0.033 (0.015-0.054)
0.060 (0.044-0.075)
SAV
2
0.019 (0.016-0.023)
0.027 (0.024-0.030)

62
Table 4.4. Recent and historic average mercury accumulation rates m cores retrieved
from Water Conservation Areas 1, 2, and 3, Everglades National Park, and Savannas State
Reserve. Numbers in parentheses indicate the range of values found.
Sample
Region
Number of
Cores
Average Mercury Accumulation
Rate (pg m'2 y'1)
1900 Post-1985
Ratio 11
Post-1985/1900
WCA1
5
14 (5-29)
79 (45-141)
7.8 (1.6-13.3)
WCA2
3
8 (4-12)
59 (35-95)
8.7 (3.9-13.9)
WCA3
3
10 (7-11)
39 (28-55)
4.0 (3.0-4.9)
ENP
5
14 (2-28)
40 (23-57)
5.9 (1.6-19.1)
SAV
2
10 (10)
34 (31-37)
3.4 (3.0-3.8)
The ratio given represents the average of the ratios for the different cores m each area,
rather than the ratio of the average for each area.
Mercury accumulation rates around the turn of the 20th century ranged from 2 to
29 pg m"2 y'1 for all cores, with apparent increasing trends beginning mid-century (1930-
1960). Post-1985 mercury accumulation rates were an average of 6.3 (1.6-19 1) times
higher than 1900 rates. Temporal changes in average mercury accumulation rates
progressed geographically from a 5.9 and 4.0 times increase in ENP and WCA1 (post-
1985/1900), to a 7.8 and a 8.7 times increase m WCA1 and WCA2 (Table 4.4). Average
mercury accumulation rates for SAV cores increased 3.4 times (post-1985/1900). The
trend of larger ratios to the north (WCA1,WCA2) with smaller ratios to the south (WCA3,
ENP) suggests at least three possible explanations:
1) some northern source of mercury m overland sheetflow;
2) non-uniform atmospheric deposition of mercury with more deposition
in northern regions;
3) non-uniform, post-depositional mobility of mercury in soils
(i.e. varying retention of mercury in different soil types) with
mercury retention decreasmg spatially from WCA1 south to
ENP.

Depth in Core (cm)
63
Water Conservation Area 1
Unsup.^'Opb (pCi/g)
0 3 6 9 12
U7Cs (pC./g)
0 2 4 6 8
Depth in Core (cm)
-45 -30 -15 (
Core 01
Bulk Density (g/cmA)
0.00 0.06 0.' 2 0.18
Sed.Rt (g cm 7y ')
0.00 0.05 0.10
Total Hg (ng/g)
0 25 50 75 100
Tot.Hg Acc.Rt. (ug m ')
0 25 50 75 100
Figure 4.3. Sediment paleostratigraphy for Water Conservation Area 1—Core 01

Depth in Core (cm)
64
Water
Unsupp.210pb (pCi/g)
03691215
-5
-10
â– 
y
i
-15
i
-20
-25
I
\
-30
Conservation Area 1
137Cs (pCi/g)
0 3 6 9 12 15
- Core 35
Bulk Density (g/cm^)
0.0 0.1 0.2
Depth in Core (cm) Sed.Rt. (g cm 2y 1)
-30 -20 -10 0 0.00 0.03 0.06 0.09
2000
c 1970
2000
1970
r
o
S 1940
1940
/
Q.
0)
^ 1910
1910
. .
1 1880
1880
r
1850
1850
i
Total Hg (ng/'g) Tot.Hg Acc.Rt. (ug m 7y 1)
0 50 100150200 0 20 40 60 80
2000
c 1970
o
O 1940
a;
Q
O 1910
> 1880
1850
Detection limit
Figure 4.4. Sediment paleostratigraphy for Water Conservation Area 1—Core 35

Depth in Core (cm)
65
Water Conservation Area
Unsupp.2 1 Opb (pCi/g) '^^Cs(pCi/g)
Core 37
Bulk Density (g/cm-'
0 10 20 0 3 6 9 0.00 0.05 0.10
Total Hg (ng/g)
0 400 800
Tot.Hg Acc.Rt. (ug m ')
0 100 200
Figure 4.5. Sediment paleostratigraphy for Water Conservation Area 1-Core 37

Depth in Core (cm)
66
Water Conservation Area 1 - Core 38
Unsuoo '' ^DS (pCi/g)
0 5
20
'-7Cs (pCi/g)
Bulk Density (g/cm~)
Depth in Core (cm)
-20-15-10-5 0
Sed.Rt. (g cm ?)
Total Hg (ng/g)
0 150 300
Tot.Hg Acc.Rt (ug m 2„-1
y J
Figure 4 6. Sediment paleostratigraphy for Water Conservation Area l~Core 38

Depth in Core (cm)
67
Water Conservation Area
UnsuppA'^Pb (pCI/g)
0 4 8 12 16
1 J7Cs (pCi/g)
0 2 4 6 8
— Core 40
Bulk Oensity (g/crrA)
0.00 0.05 0.10 0.15
Depth in Core (cm)
-20-15-10-5 0
Sed.Rt. (g cm ')
0.00 0.02 0.04
Total Hg (ng/g)
0 200 400
Tot.Hg Acc.Rt. (ug m 7y_1)
0 50 100
Figure 4.7. Sediment paleostratigraphy for Water Conservation Area 1-Core 40

Depth in Core (cm)
68
Water Conservation Area 2
Unsupp -'®Pb (pCi/g)
0 6 12 15 24
' 37Cs (pCi/g)
Core 25
Bulk Density (g/crrA)
0.1 0.2
Depth in Core (cm)
-25 -15 -5
Sed.Rt (g cm 1)
0.00 0.02 0.04
Total Hg (ng/g)
Tot.Hg Acc.Rt. (ua m ^y-')
0 200 400 600 0 50 100 150 200
Figure 4 8 Sediment paleostratigraphy for Water Conservation Area 2-Core 25

Depth in Core (cm)
69
Water Conservation Area 2 - Core 26
Unsupp.210Pb (pCi/g) 1 37Cs (pCi/g) Bulk Density (g/cm-5)
0 3 6 9 12 0 123
0.0 0.1 0.2 0.3
Depth in Core (cm)
-20-15-10-5 0
Sed.Rt. (g cm 2y ')
0.00 0.04 0.08
Total Hg (ng/g)
0 50 100150200
Tot.Hg. Acc.Rt. (ug m 2y ')
0 20 40 60
Figure 4.9. Sediment paleostratigraphy for Water Conservation Area 2-Core 26

Depth in Core (cm)
70
Water Conservation Area 2 — Core 29
Unsup7(pCi/g) (pCi/g) Bulk Density (g/crrA)
0 2 4 6 8 10
0 2 4 6 8
0.0 0.1 0.2 0.3
Depth In Core (cm)
-35-25-15 -5
Sed.Rt. (g cm ^y-')
0.00 0.03 0.06
Total Hg (ng/g)
0 50 100150200
Tot.Hg Acc.Rt. (ug m ^y 1)
Figure 4.10. Sediment paleo stratigraphy for Water Conservation Area 2—Core 29

Depth in Core (cm)
71
Water Conservation Area 3 - Core 13
Unsupp.7,<7Pb (pCi/g)
0 4 8 12 16
U7Cs (pCi/g)
0 2 4
Bulk Oensity (g/crrA)
0.0 0.1 0.2 0.3
0
—^
0
0
\
)
/
-5
-5
-5
/
c
\
T
-10
/
/
-10
1
)
-10
/
I
•
i
•
/
•
\
-15
â– 1
-15
i
-15
\
/
J
-20
-20
-20
Depth in Core (cm)
-20-15-10-5 0
Sed.Rt. (g cm ^y ')
0.00 0.03 0.06
Totol Hg (ng/g)
0 50 100150200
Tot.Hg Acc.Rt. (ug m ^y 3
0 20 40 60
Figure 4.11. Sediment paleostratigraphy for Water Conservation Area 3-Core 13

Depth in Core (cm)
72
Water Conservation Area 3 — Core 1 5
UnsuppA'^Pb (pCi/g)
0 5 10 15 20
137Cs (pCi/g)
0 2 4 6 8
Bulk Density (g/crrA)
0.0 0.1 0.2
Depth in Core (cm)
-20-15-10-5 0
Sed.Rt. (g cm 7y 1)
0.00 0.02 0.04
Total Hg (ng/g) Tot.Hg Acc.Rt (ug m 7y-1)
0 60 120180240 0 20 40 60
Figure 4.12. Sediment paleostratigraphy for Water Conservation Area 3--Core 15

Depth in Core
73
0
-10
-15
-20
Water Conservation Area 3
llnsupD.^1®Pb (pCi/g) 1 ^7Cs (pCi/g)
0 3 6 9 12
0 3 6 9 12
Core 1 9
Bulk Density (g/crrA)
0.0 0.1 0.2
Depth in Core (cm)
-20-15-10-5 0
Sed.Rt. (g cm ^y— 1)
0.0 0.1 0.2 0.3
Total Hg (ng/g)
0 50 100150200
Tot.Hg Acc.Rt. (ug m ^y ^)
0 30 60 90
Figure 4.13. Sediment paleostratigraphy for Water Conservation Area 3-Core 19

Depth in Core (cm)
74
Taylor Slough: Core 1( ), Core 2 (• -)
UnsupT'^b (pCi/g)
0 3 6 9 12
,J7Cs (pCI/g)
0.0 0.5 1.0
cm')
Depth in Core (cm)
-25 -15 -5
Bulk Density (g/
0.0 0.3 0.6 0.9
Sed.Rt. (g cm 7y
0.0 0.1 0.2
Total Hg (ng/g)
0 25 50 75 100
Tot.Hg Acc.Rt. (ug m ^y—')
0 20 40 60 80
Figure 4.14. Sediment paleostratigraphy for Everglades National Park-Taylor Slough

Depth in Core (cm)
75
Everglades National Park - Core 7
UnsupP'^Pb (pCI/g)
' ^7Cs (pCi/g)
0.0 0.5 1.0 1.5 2.0
8ulk Density (g/crrP)
0.0 0.2 0.4
Depth m Cere (cm)
-20-15-10-5 0
Sed.Rt. (g cm ^y 1)
0.00 0.05 0.10
Total Hg (ng/g)
0 30 60 90 120
Tot.Hg Acc.Rt. (ug m ^y —
Figure 4.15. Sediment paleostratigraphy for Everglades National Park-Core 7

Depth in Core (cm)
76
E\erglades National Pari< — Core 09
Unsupp 21 °Pb (pCi/g)
0 3 6 9 12
n7Cs (pCi/g)
0.0 0.5 1.0 1.5 2.0
Bulk Density (g/cm2)
0.0 0.1 0.2
Depth in Core (cm)
-20-15-10-5 0
Sed.Rt. (g cm 2y~ ')
0.00 0.03 0.06
Total Hg (ng/g)
0 50 100150200
Tot.Hg Acc.Rt. (ug m 2y“^)
0 20 40 60
Figure 4.16. Sediment paleostratigraphy for Everglades National Park—Core 9

Depth in Core (cm)
77
Everalaaes National Park - Core 1 1
Unsupp/' ®Pb (pCi/g)
0 <1 3 12
1 -57Cs (pCi/g)
0 2 4
Bulk Density (g/cm^)
0.0 0.1 0.2 0.3
0
\
0
0
I
\
l
j
-5
-5
X
-5
/ â– 
-10
1
/
-10
-10
\
{
Í
\
-15
7
-15
â–  [
-15
\
-20
-20
-20
Depth In Core (cm)
-20-15-10-5 0
Sed.Rt. (q cm ')
0.00 0.03 0.06
Totol Hg (ng/g) Tot.Hg Acc.Rt. (ug m_^y ')
0 50 100 150 200 0 20 40 60
Figure 4 17. Sediment paleostratigraphy for Everglades National Park-Core 11

Depth in Core (cm)
78
Evergiodes National Park
JPb (pC./q)
3 12
,37Cs (pC./g)
0 2 4 6 3
Depth in Core (cm)
-20-15-10-5 0
Core 12
Bulk Density (g/cm“1
0.0 o.; 0.2 0.3
0
r
-5 â– 
10 I-
-15 â– 
)
\ \
-20
Sed.Rt. (g cm ')
0.00 0.03 0.06
Totol Hg (ng/g)
0 50 100150200
Tot.Hg Acc.Rt. (ug m ')
0 30 60 90
Figure 4.18. Sediment paleostratigraphy for Everglades National Park—Core 12

Depth in Core (cm)
79
Savannas State Reserve — Core 48
Unsupp. (pCi/g) ^^CstpCi/g) Bulk Density (g/cm^)
0 2 4 6 8
0 5 10 15 20
0.0 0.2 0.4
0
0
0
;
r
r
-5
\
/
-5
.
/
/
-5
\
-10
[
"\
J
-10
f
\
I
-10
(
-15
/
-15
-15
\
-20
-20
-20
Depth in Core (cm)
-15 -10 -5 0
Sed.Rt. (g cm ^y ')
0.00 0.05 0.10
Total Hg (ng/g)
0 50 100 150
Tot.Hg Acc.Rt (ug m ^y ')
0 20 40 60
Figure 4.19. Sediment paleostratigraphy for Savannas State Reserve-Core 48

Depth in Core (cm)
80
Savannas State Reserve — Core 49
Unsupp. (pCi/g) 1^Cs (pCi/g) Bulk Density (g/cm^)
0 2 4 6 8 0.0 0.2 0.4
0 5 10 15 20
Depth in Core (cm)
-15 -10 -5 0
Sed.Rt. (g cm ')
0.00 0.05 0.10
c
o
¡ñ
O
a
o
o
o
o
2000
1970
1940
1910
1880
1850
Totol Hg (ng/g)
0 50 100 150
Tot.Hg Acc.Rt. (ug m ^y ')
0 20 40 60
¡ ^
2000
.
I â– 
1970
I X
1
1
1940
/ .
!
-1 X
1910
Í
i J
1880
1
1 *
1850
Detection limit
Figure 4.20. Sediment paleostratigraphy for Savannas State Reserve-Core 49

81
Agricultural runoff is delivered to the Everglades from agricultural land to the
north (Reddy et ah, 1991). Agricultural practices in the Everglades Agricultural Area
have increased erosion and oxidation of organic soils (Blake, 1980). Historically,
mercurial fungicides were used in this region to enhance agricultural production. Many
agricultural practices (i.e. repeated drying and flooding, and the application of mercurial
fungicides), and the resulting erosion and oxidation of soils, can facilitate the transport
of mercury (Lodenius et ah, 1987) from agricultural land to surrounding areas. Although
there is no basis, from existmg data, to quantify the relative contribution of agricultural
practices in the EAA to mercury accumulation m the Everglades system, previous studies
have demonstrated the deleterious effects of agricultural runoff on the Everglades wetland
system (Horvath et ah, 1972; Richardson et ah, 1990; Reddy et ah, 1991). The northern
portions of the Everglades (WCA1 and WCA2) likely receive atmospheric and drainage
mputs of mercury from the Everglades Agricultural Area (EAA) via mill production and
the burning of crop material (121 Kg y'1; 1981 to 1990)(KBN Engineering and Applied
Science, 1992) and the irrigation of agricultural runoff waters (Richardson et ah, 1990;
Reddy et ah, 1991).
A Florida mercury emissions survey identified four primary anthropogenic sources
of mercury in 1990, including MSW (municipal solid waste) combustion (14.6%), medical
waste incineration (14.0%), paint application (11.1%), and electricity production (10.7%)
(KBN Engineering and Applied Sciences, Inc., 1992). Natural processes contnbuted
38.9% of the total 1990 mercury emissions. The survey did not identify the relative
contributions of these sources to mercury deposition m the state. Globally, studies have

82
identified significant discharges of mercury from manufacturing and mcmeration activities
(Fukuzaki et ah, 1986), and regional gradients of mercury accumulation from point source
emissions have been identified (Nater and Gngal, 1992). Smce there is little direct
quantitative information regarding mercury deposition in Florida, at this time, spatial
differences in mercury emissions must be used as a surrogate measure of mercury
deposition.
Palm Beach, Broward, and Dade counties follow the southeast coastlme of Florida
such that Palm Beach county resides east of WCA1, Broward county resides east of
WCA2 and northern WCA3, while Dade county resides east of southern WCA3 and ENP.
The estimated total 1990 mercury emissions for Palm Beach, Broward, and Dade counties
were 1512, 1995, and 4614 Kg y'1, respectively (KBN Engineering and Applied Sciences,
1992).
To predict non-uniform mercury deposition resultmg from local anthropogenic
activities, one would expect to find the greatest mercury enrichment to occur m southern
regions of the Florida Everglades (i.e. ENP and WCA3). The reverse trend is suggested
by the sediment data (Table 4.4). Alternatively, it must be recognized that mean
estimates of mercury accumulation rates for the hydrologic basms (WCA's and ENP)
resulted from soil cores with considerable between-site variability. Further, the central
regions of Water Conservation Areas 1 and 2 are closer in proximity to the Atlantic coast
of peninsular Florida (20 and 30 km, respectively), than those of Water Conservation Area
3 and Everglades National Park (45 and 60 km, respectively).
Variable retention of mercury between organic soils and marl sediments may
influence the ambient concentrations m these substrates. The three Water Conservation

83
Areas have organic-rich soils, with a total organic carbon content (g g'1) between 40 and
50%. Sediment in ENP represents an array of marl and organic deposits, with a total
organic carbon content of 10-20%. If mercury retention were variable between organic
and marl deposits, and mercury inputs were uniformly distributed, then post-depositional
mercury migration would alter the apparent mercury accumulation rates in the mineral
sediment of ENP (Barrow and Cox, 1992a, 1992b).
Mercury retention by organic and mmeral substrates, can be compared by
exammmg average mercury accumulation rates among the Everglades regions (WCAs and
ENP). For this purpose, let us assume:
1) atmospheric mercury deposition serves as the primary mercury
mput to WCA3 and ENP soil,
2) mercury deposition is uniform over WCA3 and ENP soil,
and
3) mercury retention by organic soil exceeds mercury
retention by mmeral sediment.
One would predict that mercury accumulation rates in ENP sediment would be less than
contemporaneous rates in WCA3 soil. Assumptions 1 and 2 likely pertain to pre¬
development (1900) mercury accumulation rates m ENP and the WCAs. The average
1900 mercury accumulation rate for ENP cores (14 pg nT2 yr"‘) is similar to average
mercury accumulation rates for WCA1, WCA2, and WCA3 cores (14, 8, 10 pg m'2 yr"1,
respectively). Further, average post-1985 accumulation rates are similar for ENP and
WCA3 cores (40 and 39 pg m"2 yr"1, respectively). Enhanced mobility of mercury, that

84
may occur m marl sediment (Barrow and Cox, 1992a, 1992b), is not demonstrated by the
dated cores examined in this study.
The data suggest that mercury retention is not influenced by variability m soil
composition Further, increases m accumulation rate ratios for Everglades regions moving
from south (ENP) to north (WCAl)(Table 4.4) suggest a regional factor(s) that results in
more pronounced mercury accumulation in the northern Everglades (WCA1 and WCA2)
as compared to WCA3 and ENP.
Mercury accumulation rates mcrease gradually smce the turn of the century and
mcrease more distinctly by mid-century (1930-1960). Twelve of the eighteen mercury
accumulation profiles exhibit dramatic mcreases beginning in the 1970's and 1980's
(Figures 4.3-4.20). Some sites show increased mercury accumulation during the last two
decades with constant sediment accumulation rates (WCAP37, WCA2:26, ENP:07) or,
alternatively, with uniform mercury concentration (WCAP01, WCA2:29, WCA319,
ENP 09, ENP:11). The covariance of sediment component mputs demonstrated m these
cores illustrates the limitations of characterizing mercury deposition using mercury
concentration profiles alone.
Gradual post-1900 mcreases in mercury accumulation rates match trends found m
other systems. These mcreases are probably related to global atmospheric mcreases
resulting from European and American industrialization smce the turn of the 20th century.
Mid-century mcreases in accumulation rate are likely related m part to regional
urbanization and agriculture m south Florida (Blake, 1980).
Urban development along the southeast coast of peninsular Florida expanded
dramatically smce the 1940's (Blake, 1980). Municipal solid waste mcmeration began in

85
1951 on Florida's southeast coast. Between 1951 and 1972, the construction of eleven
incinerator facilities was completed in Dade and Broward counties. All of these facilities
were shut down by 1979 as a consequence of their inefficient emission controls. The
cumulative mercury emissions from waste incineration in southeast Florida mcreased from
955 Kg y'1 in 1951 to a maximum of 1870 Kg y'1 in 1973 with a decrease to 0 Kg y'1 in
1979 (KBN Engineering and Applied Sciences, 1992). Mercury emissions from
modernized incineration facilities, constructed since 1983, have mcreased due to increased
facility throughput (tons per year)(Table 4.5).
Table 4.5 Mercury emission estimates for municipal solid waste mcmeration in southeast
Florida (Palm Beach, Broward, and Dade counties)(KBN Engineering and Applied
Sciences, Inc., 1992).
Year
Low
(Kg y1)
Average
(Kg y1)
High
(Kg y'1)
1982
0
0
0
1983
504
575
671
1984
457
522
610
1985
388
443
518
1986
272
310
362
1987
389
445
519
1988
342
391
456
1989
447
511
597
1990
700
834
1105
1991
1225
1471
1776

86
Statewide mercury emissions in Florida, from the electric utilities industry, were
estimated to have increased 51%, from 2,062 Kg y'1 between 1981-82 to 3,111 Kg y'1
between 1989-90 (KBN Engineering and Applied Science, 1992). Southeast Florida
mercury emissions, m 1990, averaged 79 (8-203) Kg y'1 from electricity production, 835
and 1,820 Kg y’1 from MSW and medical waste incineration, respectively. Despite
improved emission controls, increasmg mercury emissions smce 1980, result from the
demands (utilities and waste control) imposed by rapidly developmg regions in the state
Error Analysis of Sediment Datmg
Errors associated with the statistical fluctuations of nuclear decay and with the
application of this uncertainty in the CRS datmg model were determmed for three
different sites (WCAF01, ENP:11, and SAV:49). Uncertainty for all other sites was
assumed to be of similar magnitude.
"Error bars" (one standard deviation on either side of the data pomt) for activity
of radio chemicals are shown in Figures 4.21-4.23. Because the predicted standard
deviation for random processes, such as gamma disintegrations, equals the square root of
the mean count, samples with a high count have a small standard deviation (as percent
of that mean). Standard deviations generally ranged from 3-6% of the mean for the
higher activity deposits to 6-30% for deeper core sections. Errors m the activity of
unsupported 210Pb were larger (generally 3-12% and 12-54%, respectively), because they
are computed as the difference of two uncertain activities.
Datmg uncertainty increased with age of the sediment (Figures 4.24-4.26). Monte
Carlo simulations (Palisade Corp., 1990) were used to calculate 500 different 210Pb

Depth in Core (cm) Depth in Core (cm)
87
Activity (pCi/g)
0 3 6 9 12
Figure 4.21. Error associated with radionuclide determinations for WCA1:01.

Depth in Core (cm) Depth in Core
88
Activity (pCi/g)
E
o
0 12 3 4
Everglades National Park - Core 1 1
Figure 4 22. Error associated with radionuclide determinations for ENP ll.

Depth in Core (cm) Depth in Core
89
Activity (pCi/g)
0 5 10 15 20
Savannas State Reserve — Core 49
Figure 4 23. Error associated with radionuclide determinations for SAV:49

Year of Deposition Year of Deposition
90
Depth in Core (cm)
-45 -35 -25 -15 -5
0.0 0.1 0.2 0.3
Figure 4.24. Dating uncertainty associated with WCA1:01.

Year of Deposition Year of Deposition
91
Depth in Core (cm)
-20 -15 -10 -5 0
0.00 0.02 0.04 0.06 0.08
Figure 4.25. Dating uncertainty associated with ENP:11.

Year of Deposition Year of Deposition
92
Depth in Core (cm)
-15 -10 -5 0
Savannas State Reserve — Core 49
Figure 4 26. Dating uncertainty associated with SAV:49.

93
profiles with the same number of dates and sedimentation rates for every core section.
Ninety-five percent confidence intervals ranged from ± 1 year in deposits laid down 10
years before present, ± 2 years at 20 years, ± 3 years at 40 years, ±5-10 years at 90
years, and ± 10-25 years at 120 years before present. These ranges corresponded with
error estimates reported by Bmford (1990) for Florida lake cores analyzed with alpha
spectrometry Large dating errors m the bottom sections of the cores made 210Pb-dates
unreliable for sediments older than 120 years. The activity of 210Pb m these old sediments
is low due to its half-life of 22.3 years. Such low mean net-counts for 210Pb result in
large standard deviations (as percent of that mean). The large age-range for the bottom
sections of ENP:11 (Figure 4.25) resulted mainly from a conscious attempt to count
samples quickly. Longer counting times reduce error.
Monte Carlo confidence mtervals (95%), expressed as a percent of the mean, for
the rate of sedimentation also mcreased with the age of the sediment. The ranges were
± 7-10% in sediments 10 years of age, ± 9-12% at 40 years, ± 15-52% at 90 years, and
± 30-90% at 120 years old. Confidence mtervals for sedimentation rates in older deposits
were as high as ± 168%. The top sections of core WCAL01 produced a wider range of
sedimentation rates (± 16-20%) compared with other cores (± 5-7%). Sedimentation rates
were calculated as a function of the activity of unsupported 210Pb. Because the low mean
net-counts for 2l0Pb m the top sections of WCA1:01 resulted m a high standard deviations
(as percent of that mean), rate calculations based on that low 210Pb activity showed more
uncertainty. The sedimentation rate at the top of that core was 0.24 ± 0 04 g cm'2 y1.

94
Sediment Mercury Concentrations
Comparison of Recent and Historic Mercury Concentrations
Recent (post-1985) and pre-development (1900) mercury concentrations were
compared to show relative changes in mercury abundance (Table 4.6). Recent mercury
concentrations were calculated as the weighted average (i.e. normalized for variations in
percent solids) of the top four centimeters of sediment. The weighted average mercury
concentration corrects for depth-specific changes in bulk density. The 1900 stratum was
determined, for all cores in a region. The 1900 stratum is defmed as those core sections
that encompass the "1900" period identified in dated cores from that region. The
weighted average mercury concentrations were determined for that stratum from each
core. For example, dated profiles from three cores in WCA3 suggested that sediment
accumulated around 1900 was between 11 and 17 centimeters beneath the sediment
surface. Therefore, the 1900 stratum for all cores from WCA3 was determined as the
weighted average concentration for the core intervals between 11 and 17 centimeters.
Fifty one cores were used to compare recent and pre-development (1900) mercury
concentrations (Table 4.6). Enrichment factors were also determined for cores retrieved
from the Stormwater Treatment Areas (STAs) of the EAA. Forty four of these cores
showed mcreased mercury concentrations m surface sediment. Four of the six sites with
lower mercury concentrations m surface sediment are m close proximity to canals
(ENP 01, WCA3:32, WCA3:33, and WCA3:34)(Figure 3.1), while the remaining cores

95
Table 4.6. Comparison of total mercury concentrations in recent and historic soils.
Sediment Core ID#
[Hg], ng/g
[Hg], ng/g
[Hg], ng/g
Enrichment
RECENT
HISTORIC
(Difference)
Factor
(0-4 cm)
(11-17 cm)
ENP: 01
20
32
-12
-0.4
ENP: 02
98
79
19
0.2
ENP: 04
38
28
10
0.4
ENP: 05
87
83
4
0.0
ENP: 06
84
16
68
4.3
ENP: 07
65
38
27
0.7
ENP: 08
140
89
51
0.6
ENP: 09
74
49
25
0.5
ENP: 10
45
25
20
0.8
ENP: 11
39
31
8
0.3
ENP: 12
99
24
75
3.1
ENRTS1
52
46
6
0.1
ENRTS2
35
27
8
0.3
WCA3:13
77
28
49
1.8
WCA3:14
263
96
167
1.7
WCA3:15
81
49
32
0.7
WCA3:16
112
54
58
1.1
WCA3:17
329
100
229
2.3
WCA318
177
85
92
1.1
WCA3:19
28
101
-73
-0.7
WCA3:20
17
10
7
0.7
WCA3:21
64
19
45
2.4
WCA3:22
75
27
48
1.8
WCA3:24
115
88
27
0.3
WCA3:01
86
57
29
0.5
WCA3:02
87
24
63
2.6
WCA3:32
11
28
-17
-0.6
WCA3:33
10
44
-34
-0.8
WCA3:34
28
63
-35
-0.6
WCA3:C123
120
60
60
1.0
WCA2:25
390
42
348
8.3
WCA2:26
154
36
118
3.3
WCA2:27
138
41
97
2.4
WCA2:28
71
50
21
0.4
WCA2:29
64
73
-9
-0.1
WCA2:30
112
38
74
1.9
WCA1:35
111
98
13
0.1
WCA1:36
244
63
181
2.9
WCA1:37
479
45
434
9.6
WCA1:38
320
116
204
1.8
WCA1:39
411
111
300
2.7

96
Table 4,6. (cont'd)
Sediment Core ID#
[Hg], ng/g
RECENT
(0-4 cm)
[Hg], ng/g
HISTORIC
(11-17 cm)
[Hg], ng/g
(Difference)
Enrichment
Factor
WCAT40
244
41
203
5.0
WCA141
147
135
12
0.1
WCA1:42
183
46
137
3.0
WCA1:01
45
76
-31
-0.4
STA: 43-47
58
55
3
0.1
SAV: 48
118
34
84
2.5
SAV: 49
120
64
56
0.9
SAV: 53
57
27
30
1.1
OKE: 56
108
70
38
0.5
OKE: 57
96
78
18
0.2
OKE; 58
91
71
20
0.3
AVERAGE (OKE)
98
73
25
0.4
AVERAGE (SAV)
98
42
57
1.5
AVERAGE (STA)
58
55
3
0.1
AVERAGE (WCA1)
182
61
132
2.2
AVERAGE (WCA2)
155
47
108
2.7
AVERAGE (WCA3)
99
55
44
0.9
AVERAGE (ENP)
67
44
24
0.8
AVERAGE (all cores)
115
53
63
1.3
Enrichment Factor (EF) = ([Fig] recent - [Hg] historic)/([Hg] historic)

97
exhibit dramatic mcreases m sediment accumulation rate m recent years (WCA1:01,
WCA3:19)(Figures 4.3 and 4 13). The average mercury concentration m surface sediment
(0-4 cm) was 121 ng g'1 (n=51, 10-479 ng g'1). The average concentration in deeper
(1900) sediment was 56 ng g'1 (10-135 ng g1). The average difference between surface
and deep sediment was 66 ng g'1 for all cores (Table 4.6). Surface sediment had an
average of 2.4 (0.2-10.6) times more mercury than deep sediment.
Sediment enrichment factors (EF) have been used by other researchers to relate
present metal concentrations to histone background concentrations (Meger, 1986; Rada
et ah, 1987; Henning et ah, 1989). The enrichment factor is calculated as the change in
metal concentration divided by the background level.
Enrichment Factor (EF) = ([Hg]rcccnt - [Hg]background)/[Hg]background
The average enrichment factor for all sites was 1.5 (-0.8-9.6)(n=51), i.e. a 150% overall
average mcrease in mercury concentrations smee the turn of the century. The ennehment
factors determined for the Everglades, Okefenokee, and Savannas State Reserve wetlands
are compared to previous lake studies (Table 4.7).
The average EF for the Everglades, Okefenokee, and Savannas falls within the
ranges identified in previous aquatic system studies. Variability of the EF m these
wetlands likely results from spatial variation in soil and habitat type, hydrologic
variability found in wetlands, and the large number of cores taken (n=45). However, the

98
Table 4.7. Comparison of mercury enrichment factors (EF) in sediment core profiles.
Enrichment Factor Number
of Cores
Location
References
0.8 - 2.8
Wisconsm lakes
Rada et aL (1987)
1.9 - 2.5
3
Minnesota lakes
Meger (1986)
0.8 - 4.5
5
Minnesota lakes
Henning et af (1989)
SO
ON
00
o'
1
45
Everglades
This Study
0.9 - 2.5
3
Savannas
This Study
0.2 - 0.5
3
Okefenokee
This Study
1.4
51
Average
This Study
similarity among EF's for these wetlands and lakes indicate similarities in: 1) trends of
mercury accumulation to these systems, and/or 2) the physicochemical processes affecting
mercury in lake and wetland sediments.
The enrichment factor is confounded by varying sediment accumulation over time
(Henning et af, 1989). In some cases, temporal mcreases m sediment accumulation will
indicate mercury depletion in surface sediment. Accordmgly, the enrichment factor
cannot assess changes in mercury mput. However, the EF can be used to characterize
temporal changes m mercury abundance at the sediment-water interface where it is most
available for biotransformation. As such, the enrichment factor provides a measure of
spatial or temporal difference in mercury bioavailability.
Identification of the bioavailable component of sediment mercury is still poorly
understood. Selective extraction procedures have been suggested as a surrogate for the
direct determination of bioavailable mercury (Duddndge and Wainwright, 1991). These

99
selective extraction procedures indicate that less than ten percent of the total metal
concentrations are solubilized by extractants commonly used to predict plant mercury
uptake from soils. The bioavailability of mercury is directly related to the total organic
material content of sediment (Langston, 1982). Bioavailability, characterized by
comparing tissue mercury content with sediment mercury:organic ratios, is greater in
mm eral sediment with low mercury concentrations than m organic sediment with high
mercury concentrations. An earlier study of mercury in Everglades sediment suggested
that mercury was strongly associated with organic matter and sulfide complexes (Lmdberg
and Harnss, 1974). Smce average mercury concentrations have increased 150% m the
past 100 years, the bioavailable mercury fraction has likely increased as much as 2.5
times smce the turn of the century.
Post-Depositional Mobility of Mercury
Forty four sediment cores showed elevated mercury concentrations in surface
layers (Table 4.6). Numerous studies suggested that mercury does not migrate readily in
sediment (Rogers and MacFarlane, 1978; Wallace et ah, 1982; Lodenius et ah, 1987;
Henning et ah, 1989; Lodenius, 1990, Wmfrey and Rudd, 1990; Schuster, 1991; Barrow
and Cox, 1992a, 1992b; Bryan and Langston, 1992; Gobiel and Cossa, 1993 in press).
Henning et ah (1989) attempted to identify post-depositional mobility of mercury
in a 14 month core mcubation study. Spiked sediment (1200 ng Hg2+ g'1) was transferred
quantitatively into sediment core tubes or was transferred as the top 2 centimeters into
unspiked cores. Cores were mcubated, m the dark, under aerobic and anaerobic

100
conditions. No evidence of mercury transport was observed m the sediment, even with
extensive tunneling by benthic organisms. No detectable mercury was found m sediment
porewater after the mcubation period.
Core studies of mercury concentrations in sediment and interstitial waters in the
Laurentian Trough of the lower St. Lawrence Estuary demonstrated that interstitial water
of Laurentian Trough sediments is enriched in mercury relative to the overlying bottom
water (Gobiel and Cossa, 1993 in press). Although mercury concentrations in porewater
were higher than those found in overlying water, these researchers concluded that
"redistribution of remobilized mercury subsequent to deposition could not explain more
than a small proportion of the 100-400% variations with depth of the mercury
concentrations in the sediment observed at all stations". Further, they concluded that
sediment mercury concentration profiles in the Laurentian Trough primarily reflected the
temporal changes m mercury mput.
Most soil mercury is adsorbed on sites with high binding energy, primarily
strongly bonded to humic matter (Lodenius, 1990). Mercury retention by organic matter
is greater than mineral substrates. Mineral soils more readily released mercury with
increasing salinity, or decreasmg pH (Barrow and Cox, 1992a), than organic-rich soils
(Barrow and Cox, 1992b).
Lodenius et af (1987) demonstrated that mercury desorption from peat soils was
not affected by changing salinity, fertilization, or sterilization. However, they showed that
complete drying and wetting of peat soils may alter the physical properties of the soil.
Deep crackmg of dried peat soils facilitates a mechanism by which mercury can be

101
released from the soil matrix into overlymg water m particulate form. Deep crackmg was
not evident during sampling for this study.
Rogers and McFarlane (1978) showed that 20% of mercury applied to soil was
volatilized rapidly, but the remaining 80% was sequestered by the soil and rendered
unavailable to microbial activity. The high affinity of mercury for sulfide accounted for
the strong bindmg of mercury to soil organic matter and to the stability of HgS (Schuster,
1991). This affinity of mercury for sulfidic compounds caused mercury to be particularly
immobile m organic-rich sediment and, consequently, less available to biota. The
physical fractionation of soil organic matter (dissolved vs. adsorbed) determmed the
behavior and distribution of mercury m soils (Schuster, 1991).
The affinity of mercury for organic matter was the most important factor governing
its chemical speciation, transport, and toxicity in Controlled Experimental Ecosystems
(CEEs)(Wallace et af, 1982). Mercury was readily scavenged by particulate organic
matter and was rapidly removed from CEEs by particulate settling. This process was also
identified by Winfrey and Rudd (1990) in their study of the factors affecting mercury
methylation m low pH lakes.
Winfrey and Rudd (1990) spiked Little Rock Lake sediment with radio-labelled
mercury m laboratory experiments. Less than 1% of the spike mercury was lost as
methylmercury. This observation was further supported when Winfrey and Rudd (1990)
demonstrated that 0.36% and 0.09% of radio-labelled mercury was methylated at ambient
pH's of 5.1 and 6.1, respectively. They also showed that 0.84% of radio-labelled mercury
was methylated m anoxic sediment. Loss of 1% of the total mercury m surface sediment

102
through methylation was not sufficient to alter apparent sediment mercury concentration
profiles.
Mercury concentration profiles were examined m surface sediment (0-10 cm) of
dated Everglades sediment (Figures 4.3-4.20). Eight of the sixteen core profiles showed
distinct increases m mercury near the surface, while five cores suggested no concentration
change and three showed distmct decreases m concentration. There is no recurrent trend
in mercury concentration profiles that would suggest post-depositional mobility of
mercury in the wetland cores. However, mercury concentration profiles, compared with
sediment accumulation rate profiles, clearly show that mercury may be diluted by
dramatic mcreases m sediment accumulation. Enrichment of mercury in surface sediment
was likely due to recent accumulation and not from migration m the core profile.
Error Analysis of Mercury Determinations
Replicate measurements of total mercury were determined routmely to identify the
error associated with mercury determmations. Duplicate measurements were made for at
least one sample per set of analyses. The typical variability was ± 6 ng g'1 (± 0 to ± 18
ng g'1; n=42) for duplicate mercury measurements performed on the same day. Further,
a holdmg time study was implemented to demonstrate the viability of freezing samples
during storage. For replicate analyses performed on non-consecutive days (6-54 day
holdmg interval), the data suggest a typical range of ± 17 ng g'1 (n=80) for sediment
mercury concentrations. Variations m percent solids, between dissimilar samples, does
not appear to influence reproducibility, but sample mixing and daily variations in
instrument calibration does influence reproducibility of mercury determmations.

103
Spatial Distribution of Mercury in the Everglades
Surface mercury concentrations were highest m Water Conservation Areas 1 and
2 (Figure 4.27). All WCA1 sites yielded mercury concentrations exceedmg 100 ng g"1
(111-411 ng g'1). Four of the six sites m WCA2 exceeded 100 ng g'1 (112-390 ng g'1),
while five of fifteen sites m WCA3 exceeded 100 ng g'1 (112-329 ng g1). One core
(ENP:08) of the twelve cores (or surface grab samples) collected m the ENP had a level
of 140 ng g'1.
Eighteen of 45 sites sampled throughout the Everglades showed surface
concentrations greater than 100 ng g1. Within the subgroup of 18 sites exceedmg 100
ng g'1, seven sites had mercury concentrations over 200 ng g'1, and four sites showed
mercury levels over 300 ng g'1. Historic (1900) mercury concentrations throughout the
Everglades exceeded 100 ng g'1 at only five sites (3 WCA1 sites, 2 WCA3 sites)(Figure
4.28).
Mercury concentrations were compared for duplicate cores, taken 1 to 2 meters
apart, at three sites (ENP:TS1,2, ENP:08a,b; WCA2:26a,b). The mean variability between
corresponding samples was ± 12 ng g'1 (± 1 to ± 68, n=39).
Organic-rich sediment from the Water Conservation Areas exhibited higher overall
mercury concentrations than the mineral sediment m the southeast Everglades National
Park. Some organic sediments m the WCA's have low mercury concentrations similar to
those of mineral sediment, however, these locations were dry during our sampling period,
and are m regions of the Everglades that experience frequent, extended periods of drying.
Repeated drying and flooding have been identified as the primary mechanism for leachmg

104
Figure 4.27. Spatial distribution of mercury (ng g'1) in post-1985 (0-4 cm) sediment.

80° OCT
81°30’ 8 Io 00’
i
80° 30’
26° 30’ -
26° 00’ -
25° 30' -
25° 00' -
Figure 4.28. Spatial distribution of mercury (ng g'1) in histone (ca. 1900) sediment.

106
mercury from peat soil (Lodemus et ah, 1987). The sites exhibiting low mercury
concentrations m organic sediment correspond to regions where present-day conditions
are considerably more variable and typically drier than pre-development conditions
(SFWMD, 1992). The northern region of WCA3 (sites 20-24) and the northeast region
of ENP (sites 1-4) have an average surface concentration (0-4 cm) of 66 (17-98) ng g'1
The spatial distribution of mercury m pre-development (1900) sediment showed
that the highest concentrations of mercury occurred m WCA1 and the western half of
WCA3, with the maximum concentrations not exceedmg 135 ng g'1. It must be noted,
however, that some of the lowest mercury concentrations occurred m the western half of
WCA3 (the dry regions to the north). The recent (post-1985) mercury spatial distribution
indicated that mercury concentrations were elevated, compared to pre-development (1900)
sediment in 38 of 45 Everglades sites. Although each sampling region (ENP, WCA1,
WCA2, and WCA3) exhibited sites with recent mercury concentrations exceedmg 100 ng
g'1, no ENP site exceeded 140 ng g_1 in recent sediment. Sites exceeding 150 ng g’1 in
recent sediment included sites throughout WCA1, southern WCA2, and southwestern
WCA3.
Relationships Between Mercury Concentration and Selected Water and Sediment
Parameters
Mercury-carbon relationship in sediment. Sediment cores retrieved from WCA1,
WCA2, and WCA3 are rich m organic peat soil while those retrieved from the Taylor
Slough hydrologic basm of Everglades National Park are comprised of marl sediments.
The mean mercury concentration for marl sediment (33 ± 8 ng g1, n = 36), although not

107
statistically different from that of the organic-rich sediment, was generally less than that
for organic-rich sediment (80 ± 45 ng g'1, n = 99). Even though organic sediments
have a stronger mercury binding capacity than mineral sediment (Schuster, 1991), there
is no apparent relationship between mercury concentration and percent total organic
carbon (r2 = 0.23, n = 57) m these soils.
All total carbon determmations were compared to their corresponding total
mercury concentration. Mercury concentrations in organic-rich sediment (total carbon
>30%)( 10 to 808 ng g1) encompassed a larger range than m marl sediment (total carbon
<30%)( 10 to 110 ng g1), and spanned the entire concentration range identified for marl
sediment. The larger average mercury concentration m organic sediment than m marl
sediment is attributed to the stronger mercury binding capacity of organic sediments
(Barrow and Cox, 1992a, 1992b) or to nonuniform mercury inputs.
Mercury content m soils is determined by 1) sedimentation rate, 2) mercury mput,
and 3) physico-chemical sorption characteristics. As a result, mercury concentrations in
marl sediments may be determmed by 1) the quantity of mercury mput to a sediment
volume and 2) the ability of mercury to be retamed m the sediment.
Mercury (sediment 1-conductivitv (water) relationship. Mercury adsorption to soil
and sediment may be influenced by chloride concentration (Lodenius, 1990; Schuster,
1991). Soil mercury sorption is not influenced by pH mcreases from 4 to 6 m the
absence of chloride, however, m the presence of chloride, mercury sorption mcreases with
a pH transition from 4 to 6, with a subsequent decrease at pH>6 (Barrow and Cox, 1992a,
1992b). Natural concentrations of chloride found m precipitation (ca. 20 pM) and lake

108
water (ca. 200 pM) are not sufficient to facilitate considerable desorption of mercury from
soils (Lodemus et af, 1987).
The conductivities (pmhos cm'1) of overlymg water in this study, based on single-
event determmations, were compared to the total mercury concentrations of surficial
sediment (0-4 cm). Conductivity was used as a surrogate for chloride concentration. The
correlation between mercury concentration and conductivity (r^O.15, n=36) suggests that
there is little relationship between ambient conductivity and mercury concentration m
Everglades sediment. Conductivity measurements performed during smgle-event sampling
may not be representative of the average conductivities found at these sites, however, in
the absence of a comprehensive long-term water quality database, these data provide a
preliminary evaluation of the influence of ambient chloride concentration of sediment
mercury desorption.
Supplementary Sediment Metals Concentrations
Six supplementary metals (Cd, Cu, Cr, Ni, Pb, and Zn) were determined m
selected core sections (Number of cores=13, n=176)(Appendix). These metals have been
demonstrated to serve as tracers of anthropogenic emissions from mcmeration and
combustion activities (Thornton and Abrahams, 1984). Metal concentrations in
Everglades soils, were compared to those found m other studies, to characterize the
magnitude of trace metal contamination. Temporal changes m accumulation of these
metals were calculated to examine correspondence between changing metal accumulation
and regional anthropogenic activities.

109
Metals concentrations in Everglades soils were compared to those found in lakes
and wetlands, worldwide, and to surface soil concentrations previously determined for the
Everglades (Table 4.8). All Everglades metal concentrations fell within the ranges
established for the described lakes and wetlands, and agreed well with previously
determined Everglades data (Table 4.8).
Trace metal accumulation rates were calculated for dated cores from the
Everglades. The geochronology of average accumulation rates was determmed for each
metal (Cd, Cr, Cu, Ni, Pb, and Zn) (Figures 4.29-4.34). Average accumulation rates
remained constant from the late 1800's until the mid-1900's for all the metals, and rates
did not vary, between sites, during the first half of the 20th century. This is demonstrated
as narrow "range bars" m Figures 4.29-4 34 for deeper ("older") soils. All metals showed
trends of increasing accumulation rates beginning mid-century (1940-1960) and showed
mcreasmg variability, between sites, for post-1950 strata. Turn of the 20th century and
post-1985 average accumulation rates were calculated for all metals (Table 4 9).
Post-1985 accumulation rates for the trace metals are greater than corresponding
histone rates. Cadmium and lead showed two- and three-fold mcreases, respectively, m
accumulation rate smee 1900. Chromium and zinc accumulation rates mcreased four-fold,
while copper and nickel accumulation rates mcreased seven-fold.
Studies of trace metal inputs to Lake Windermere, England (Forstner, 1984)
identified recent accumulation rates for copper, lead, and zinc of 3.0, 12.5, and 24.4 mg
m"2 y'1, respectively. These rates were 2.5, 4.8, and 3.3 times greater, for copper, lead,
and zinc, respectively, than the corresponding natural baselme accumulation rates. These
rate mcreases are similar to those found m Everglades cores.

Table 4.8. Comparison of trace metal concentrations (mg Kg ') in Florida Everglades sediment with concentrations reported for other systems.
Location
[Cd]
[Cu]
[Cr]
[Ni]
[Pb]
[Zn]
[Fe]
References
Global Mean Sediment
0.2
33
N/A
N/A
19
95
41000
Salomons and Forstner (1984)
UK Estuaries
0.7
92
41
26
106
265
26246
Bryan and Langston (1992)
Okefenokee Swamp, Georgia
Minnie's Lake
N/A
11
15
5
19
13
1200
Casagrande and Erchull (1976)
Chesser Prairie
N/A
39
38
5
16
23
2280
Laguna Lake, Phillipines
N/A
96
N/A
N/A
34
210
34750
Vincente-Beckett et ah (1991)
Lindisfarne Bay, Tasmania
5
35
N/A
N/A
155
405
N/A
Wood et ah (1992)
Carezza Lake, Antarctica
0.4
24
103
N/A
10
44
50340
Campanella et ah (1991)
Sepetiba Bay, Brazil
N/A
4
N/A
N/A
47
438
3141
Quevauviller et ah (1992)
Florida Everglades (Surface grab sai
mples; 8/15/89-11/11/89)
Mean
0.9
12
16
7
18
24
7160
South Florida Water
Minimum
0.1
0.1
0.5
0.5
0.3
0.5
41
Management District
Maximum
3.9
82
150
57
99
170
44000
(unpublished data)
Florida Everglades (Sediment cores
from present study)
All Cores: recent (post 1985)
<3
<15
<15
<6
57
30
5142
Present Study
historic (1900)
<3
<17
<13
<6
41
10
5169
Marl: recent (post 1985)
4
<16
<14
13
77
23
9060
historic (1900)
4
<26
<13
15
68
10
9496
Organic: recent (post 1985)
<3
<15
<16
<4
51
33
3918
historic (1900)
<3
<15
<13
<4
39
15
3384
Note: Concentrations presented as "less than (<)" indicate that some samples comprising the average were below the analytical detection limit for that
metal..

Ill
Cadmium Accumulation Rate
, . “2.
(mgTn y )
01 23456789 10
Fiaure 4.29. Average cadmium accumulation rates m Everglades soil profiles.

112
Chromium Accumulation Rate
(mg -m 2 - y 1)
Figure 4.30. Average chromium accumulation rates m Everglades soil profiles.

113
Copper Accumulation Rate
(mg -m 2 â–  y 1)
0 10 20 30 40
Figure 4.31. Average copper accumulation rates m Everglades soil profiles.

114
Lead Accumulation Rate
/ -2 -1»
(mg • m ' y )
Figure 4 32. Average lead accumulation rates m Everglades soil profiles.

115
Nickel Accumulation Rate
(mg • m 2 • y 1)
Figure 4.33. Average nickel accumulation rates m Everglades soil profiles.

116
Zinc Accumulation Rate
. -2 -1.
(mg • m - y )
Figure 4.34. Average zinc accumulation rates m Everglades soil profiles.

117
Table 4.9. Historic (1900) and recent (post-1985) trace
metal accumulation rates m Florida Everglades soil.
Element
Accumulation Rate (mg m'2 y '')
Historic
Recent
(1900)
(post-1985)
Cadmium
1
2
Chromium
1
4
Copper
2
14
Nickel
1
7
Lead
10
30
Zinc
7
30
Estimates of global emissions of trace metals to the atmosphere from natural and
anthropogenic sources (Table 4.10) show the dramatic impact that human activities have
had on the trace metal cycles. Although remote areas, such as the Greenland ice cap,
have recently shown enhanced concentrations of trace metals commonly associated with
pollution (Cu, Zn, Sb, and Pb)(Zoller, 1984), most trace metal emissions impose their
greatest impact regionally (Thornton and Abrahams, 1984; Nnagu, 1990). Zoller (1984),
in a review of current research, concluded that urban regions provide a major contribution
of trace metals to the environment, mostly by high temperature processes (i .e. incineration
and combustion). He further concluded that the largest impacts of anthropogenic
emissions were on a regional scale and they occurred primarily downwind of large
industrial complexes.

Table 4.10.
Trace metal emission
to the atmosphere (106
Kg y'1
)(Nnagu, 1990).
Element
Natural
Anthropogenic
(106 Kg y1)
Total
Enrichment
Factor (%)
Cadmium
1.4
7.6
9
543%
Chromium
43.
31.
74.
72%
Copper
28.
35.
63.
125%
Lead
12.
332.
344.
2767%
Nickel
29.
52.
81.
179%
Zmc
45.
132.
177.
293%
Patterns of trace metal enrichment m Everglades soils, during the 20th century, are
similar to those found in European lake sediment (Forstner, 1984). Increasmg metal
accumulation, begmnmg between 1940 and 1960, is concurrent with the implementation
of major hydrologic control measures in the Everglades wetland system (1940's)
(SFWMD, 1992). These water control measures facilitated mid-century mcreases in
urbanization along southeast coastal Florida (Blake, 1980) which mcluded the onset of
large-scale solid waste incineration in 1951 (KBN Engineering and Applied Sciences,
1992). Because trace metal accumulation m the Everglades comcides with the history of
urban development m southeast Florida (Blake, 1980), and anthropogenic emissions have
been shown to establish regional gradients of metal contamination (Thornton and
Abrahams, 1984; Zoller, 1984), local anthropogenic emissions from mcmeration and fuel

119
combustion likely play a key role in the mcreases of trace metal accumulation in these
soils during this century. As a corollary, regional urban emissions likely contribute to
increased mercury accumulation m the Everglades durmg the latter half of the 20th
century.
Statewide mercury emissions in Florida, from the electric utilities industry, were
estimated to have mcreased 51%, from 2,062 Kg y'1 between 1981-82 to 3,111 Kg y'1
between 1989-90 (KBN Engmeermg and Applied Science, 1992). These mcreases durmg
the 1980's result from mcreased demand from population growth in recent years (Jurczyk,
1993) Southeast Florida mercury emissions, m 1990, averaged 79 (8-203) Kg y'1 from
electricity production, 835 and 1,820 Kg y1 from MSW and medical waste mcmeration,
respectively. Further study of trace metal accumulation m the environment may identify
the relative impact of global and regional activities on mercury accumulation m the
Everglades. Trace metal paleostratigraphy may also identify specific anthropogenic
activities, associated with regional urban development, that impose the greatest
contributions to the Everglades system.

CHAPTER 5
SUMMARY, CONCLUSIONS, AND RECOMMENDATIONS
Summary
The average mercury concentration of Everglades surface soil (0-4 cm) was 121
ng g'1 (10 to 411 ng g'1, n = 45) for all sites, and the average concentration for deep soil
("ca 1900") was 56 ng g'1 (10 to 135 ng g"1, n = 45). Mercury enrichment in Everglades
surface soil (150%, n = 45) was the same as that found in the Savannas State Reserve
(150%, n = 3) and was greater than that found m the Okefenokee Swamp (40%, n = 3).
For all study sites combmed, the average mercury concentration of surface soil was 140%
greater than concentrations identified m corresponding deep soil. There was no mercury
enrichment m surface soil from the Stormwater Treatment Areas (STA), which are
comprised of recently flooded, tilled soils of the Everglades Agricultural Area (EAA).
In the Everglades, mercury enrichment of surface soil was greater m WCA1 and
WCA2 (270%, for each) than m WCA3 and ENP (90% and 80%, respectively). Lesser
enrichment m WCA3 and ENP may result from pronounced wet/dry cycles.
The average mercury concentration m marl sediment (33 ng g'1) was less than that
for organic soil (80 ng g'1). However, the range of mercury concentrations m organic soil
encompassed the entire mercury concentration range that was identified m marl sediment.
There was little correlation between total mercury and carbon content (r2 = 0.23, n = 57).
120

121
Average mercury accumulation rates mcreased throughout the Everglades wetland
(WCA1, WCA2, WCA3, and ENP) smce the turn of the century. Mercury accumulation
around 1900 ranged from 2-29 (ig m'2 y'1 for all cores, while post-1985 accumulation
rates ranged from 23-141 pg m'2y'\ In the Everglades, mercury accumulation rate (post-
1985/1900) increases were larger m the northern regions (7.8 and 8.7 for WCA1 and
WCA2, respectively) than m the southern regions (4.0 and 5.9 for WCA3 and ENP,
respectively), although the range of rate increases for each region was variable. This
trend, however, may mdicate some atmospheric mercury input entering the Everglades
from the north. The average rate mcrease for the Savannas cores was 3.4 (3.0-3 8, n=3).
Trace metals concentrations (Cu, Cd, Cr, Fe, Ni, Pb, and Zn) matched findings of
previous studies in the Everglades and other systems worldwide. Increased accumulation
of cadmium, chromium, copper, nickel, lead, and zinc, occurred mid-century (1940-1960),
in concurrence with accelerated urbanization along southeast coastal Florida (Blake,
1980). Post-1985/1900 accumulation rate ratios showed two and three times mcrease for
cadmium and lead, respectively. A four-times mcrease for chromium and zmc and a
seven-times mcrease for copper and nickel were observed.
Conclusions
• Mercury enrichment m surface soil/sediment is the result of recent
mcreases of mercury mput not from migration of mercury m the core
profile.

122
• Mercury content is not related to soil carbon content, suggesting that
mercury input is non-uniformly distributed through the Everglades and/or
mercury content is dictated by bulk sediment accumulation.
• Ambient ionic composition (measured as conductivity) has little effect on
soil mercury concentrations.
• Radiochemical datmg of sediment is a viable tool for environmental studies
m Florida wetlands.
• Mercury inputs must be presented as accumulation rates to eliminate the
confounding effect of variable sediment accumulation.
• Trends for mercury accumulation m Florida wetlands match findings
reported for lakes in Minnesota, Wisconsm, and Sweden, perhaps indicating
a global process that leads to similar accumulation rates over widely
varying geographic regions.
• Mid-century increases of trace metal accumulation m the Everglades concur
with the history of accelerated urbanization along coastal southeast Florida.
• Ever-increasing demands for waste control (i.e. mcmeration) and electricity
production m southeast Florida are likely a dominant component of
mcreasmg metal accumulation rates found m Everglades cores.
Recommendations
• Matchmg radiochemically denved dates with independent age-markers is
needed. Examples of such age-markers may include charcoal horizons.

123
documentable pulse-inputs of contaminants (e g. DDT/DDT-metabolites,
trace metals, nuclear weapon fission products), and fossil records.
• Further studies should be implemented to identify the influence of coastal
urbanization on atmospheric mputs to the Everglades. Trace metal
transport studies may provide detail of the relative impact of industry, the
electric utilities, and waste control activities on contaminant mputs to the
Everglades.
• Micro- and macro-scale heterogeneity of mercury should be examined to
determine the interpretive resolution permitted by field studies m the
Everglades.
• Monitoring programs should be established to quantify atmospheric
mercury deposition and surface-flow mercury mputs to the Everglades.
These programs, which entail the measurement of mercury m air and water,
require "clean field and laboratory technique" and advanced analytical
technology (i.e. atomic fluorescence spectrophotometry).

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APPENDIX
FLORIDA WETLAND SOIL CHEMISTRY DATABASE

Table A. 1. Sediment and Mercury Analyses - Water Conservation Area 1
Solids
Bulk
Total
Inorg.
Organic
deposition
Total
Sampling
dry wt
Density
Carbon
Carbon
Carbon
Pb-210
Cs-137
period
rate
Mercury
Samóle I D.
Date
g/g
g/cmA3
%
%
%
pCi/g
pCi/g
year g/cmA2-vr
ng/g
WCAl :01:00-02
06/18/92
0.018
0.015
39.3
1.8
37.5
1.97
0.18
1992
0.232
77
WCA1:01:02-04
06/18/92
0033
0.023
40.2
4.02
0.16
1992
0.115
49
WCAl :01:04-06
06/18/92
0.069
0.042
43 3
6.32
0.13
1991
0.072
29
WCA1:01:06-08
06/18/92
0.076
0.080
44.1
0.9
43.2
5 33
0.19
1990
0.083
73
WCA1:01:08-10
06/18/92
0.063
0.064
40.4
9.08
0,20
1988
0.046
57
WCAl :01:10-12
06/18/92
0.052
0 044
44.7
9 82
0.37
1985
0.039
57
WCAl :01:12-14
06/18/92
0.046
0.042
45.5
0.2
45.3
11.51
028
1983
0.031
79
WCAl :01:14-16
06/18/92
0.097
0.118
40.7
6.97
1.43
1980
0.046
59
WCAl :01:16-18
06/18/92
0.084
0.085
41.2
873
1.02
1975
0.031
39
WCAl :01:18-20
06/18/92
0.088
0.060
43.4
5.64
1.40
1969
0.040
42
WCAl :01:20-22
06/18/92
0.103
0.078
4.86
5.10
1965
0.042
43
WCAl :01:22-24
06/18/92
0.121
0.111
5 69
5.68
1961
0.032
55
WCAl:01:24-26
06/18/92
0.161
0.139
43.8
7.00
5.04
1953
0.020
56
WCAl:01:26-28
06/18/92
0.131
0.123
4 69
2 64
1935
0.017
61
WCAl:01:28-30
06/18/92
0.103
0.102
2.42
1.10
1916
0.018
62
WCAl :01:30-32
06/18/92
0.092
0.070
50.9
1.69
0.51
1902
0.017
63
WCAl :01:32-34
06/18/92
0.077
0.079
1.19
0.08
1892
0.017
49
WCAl :01:34-36
06/18/92
0.083
0.072
0.69
0.11
1881
0.021
27
WCAl :01:36-38
06/18/92
0.086
0.061
49.9
1873
0.019
45
WCA1:01:38-40
06/18/92
0.094
0.063
1866
0.017
25
WCAl:01:40-42
06/18/92
0.087
0.074
1858
0015
27
WCAl:01:42-44
06/18/92
0.078
0.069
54.3
1846
0.012
13DL
WCAL01:44-46
06/18/92
0092
0081
0.33
0.10
13
WCAl:01:46-48
06/18/92
0.083
0.071
BDL
WCAL0L48-50
06/18/92
0.108
0.074
50.9
0.0
50.9
12
WCAl:01:50-52
06/18/92
0.083
0.076
BDL
WCAl :01:52-54
06/18/92
0.110
0.081
20
WCAl :01:54-56
06/18/92
0.094
0.085
53.3
BDL
WCAL0L56-58
06/18/92
0.111
0.090
29
WCAl :01:58-60
06/18/92
0.096
0.085
BDL
WCAl:01:60-62
06/18/92
0.102
0.201
50.1
0.0
50.1
21
WCAl:01:62-64
06/18/92
0.105
0.105
6
WCAl :01:64-66
06/18/92
0.112
0.095
27
WCAl :01:66-68
06/18/92
0 093
0.083
52.7
BDL

Table A. 1. (cont'd)
Solids
Bulk
Total
Sampling
dry wt
Density
Carbon
Sample I D.
Date
S/&
g/cmA3
%
WCAl:01:68-70
06/18/92
0.110
0.092
WCAl:01:70-72
06/18/92
0 102
0 091
WCA1:01:72-74
06/18/92
0.109
0.209
WCAl:01:74-76
06/18/92
0.112
0.101
49.4
WCAl:01:76-78
06/18/92
0,104
0.089
WCAl:01:78-80
06/18/92
0 093
0.080
54.6
WCA1:01:80-82
06/18/92
0.093
0.094
WCAl:01:82-84
06/18/92
0080
0.068
WCA1:01:84-86
06/18/92
0.095
0.075
50.5
WCAl:01:86-88
06/18/92
0.097
0.078
WCAl:01:88-90
06/18/92
0.094
0 094
WCAl:01:90-92
06/18/92
0.083
0.068
52.6
WCAl:01:92-94
06/18/92
0.082
0.077
WCA1:01:94-96
06/18/92
0.087
0.080
WCA1:01:96-98
06/18/92
0084
0.072
54.4
WCA1.01:98-100
06/18/92
0.105
0.089
WCA1:01:100-102
06/18/92
0.104
0 094
WCA1:01:102-103
06/18/92
0.113
0.068
52.3
WCA1:35:00-02
07/15/92
0.150
0.150
47.2
WCA1:35:02-04
07/15/92
0.163
0.168
47.3
WCA1:35:04-06
07/15/92
0 145
0.145
47.0
WCA1:35:06-08
07/15/92
0.131
0.126
46.9
WCA1:35:08-10
07/15/92
0.128
0.136
49.5
WCA1:35:10-12
07/15/92
0 136
0.137
47.7
WCA1:35:12-14
07/15/92
0.129
0.126
50.0
WCA1:35:14-16
07/15/92
0.127
0.125
50.4
WCA1:35:16-18
07/15/92
0 123
0.109
51.1
WCA1:35:18-20
07/15/92
0.140
0 124
49.7
WCAl:35:20-22
07/15/92
0 137
0.114
WCA1:35:22-24
07/15/92
0 132
0.133
WCA1:35:24-26
07/15/92
0 132
0 119
51.0
WCAl:35:26-28
07/15/92
0 129
0.114
WCAl:35:28-30
07/15/92
0.119
0.108
WCA1:35:30-32
07/15/92
0 115
0.112
52 3
deposition Total
Cs-137 period rate Mercury
pCi/g year g/cmA2-yr ng/g
24
BDL
49
19
26
83
65
85
35
47
43
44
60
63
93
30
39
32
0.0
47.2
9.13
1.47
1992
0.056
90
0.0
47,3
12.95
3.28
1986
0033
131
9.57
11.20
1974
0.031
116
8.01
10 87
1963
0.026
154
0.0
49 5
5,55
7.36
1952
0.026
174
4.78
3.71
1940
0.021
115
1.99
3 55
1924
0.031
105
0.0
50.4
076
1.72
1914
0.060
93
0.0
51.1
1.12
0.69
1910
0.035
76
0.0
497
1.30
0.06
1902
0024
191
065
0.05
1887
0030
90
0 82
0.80
1876
0017
80
0 39
0 66
1855
0 019
57
0.50
008
1836
0.008
58
63
0.0 523 45
Inorg. Organic
Carbon Carbon Pb-210
% % pCi/g
0.0 54.6
0.0 54.4

Table A. 1. (cont'd)
Sample I D.
WCAl :35:32-34
WCAl:35:34-36
Sampling
Date
07/15/92
07/15/92
Solids
dry wt
g/g
0 105
0.125
Bulk
Density
g/cmA3
0.106
0.120
Total
Carbon
%
WCAl :36:00-02
07/15/92
0.051
0.049
39 8
WCAl :36:02-04
07/15/92
0.063
0.062
41.9
WCAl:36:04-06
07/15/92
0.050
0.051
43.8
WCAl :36:06-08
07/15/92
0.051
0.050
45.3
WCAl :36:08-10
07/15/92
0.056
0.056
47.5
WCAl :36:10-12
07/15/92
0.065
0.065
487
WCA1:36:12-14
07/15/92
0.061
0.062
479
WCA1:36:14-16
07/15/92
0.062
0061
486
WCA1:36:16-18
07/15/92
0069
0.069
482
WCAl :36:18-20
07/15/92
0073
0.070
47 8
WCAl: 36.20-22
07/15/92
0.069
0.068
WCAl:36:22-24
07/15/92
0 068
0069
WCAl :36:24-26
07/15/92
0.068
0 066
WCAl:36:26-28
07/15/92
0.075
0.073
50 2
WCAl :36:28-30
07/15/92
0.067
0.065
WCAl :36:30-32
07/15/92
0.068
0.067
WCAl 36:32-34
07/15/92
0067
0.064
50.0
WCAl 36:34-36
07/15/92
0.059
0057
WCAl :36:36-38
07/15/92
0.063
0.061
WCAl :36:38-40
07/15/92
0.075
0,075
47.6
WCAl:36:40-42
07/15/92
0.069
0.068
50.3
WCAl :36:42-44
07/15/92
0.082
0.081
49.4
WCAl:36:44-46
07/15/92
0.081
0.080
WCAl:36:46-48
07/15/92
0.067
0.064
WCAl :36:48-50
07/15/92
0.087
0.088
WCAl :36:50-52
07/15/92
0.101
0.100
49.7
WCAl:37:00-02
07/15/92
0.009
0.009
42.2
WCAl :37:02-04
07/15/92
0.011
0.011
42 6
WCAl:37:04-06
07/15/92
0.008
0.008
43 4
WCAl :37:06-08
07/15/92
0.016
0.016
430
WCAl :37:08-10
07/15/92
0.058
0.056
45.4
WCAl 37:10-12
07/15/92
0.076
0.075
45 7
Inorg. Organic deposition Total
Carbon Carbon Pb-210 Cs-137 period rate Mercury
% % pCi/g pCi/g year g/cmA2-yr ng/g
34
2H_
0.0 398 203
0.0 41.9 277
0.0 438 212
205
169
116
76
90
0.0 48.2 96
63
57
36
58
33
58
36
36
42
16
0.0 47.6 13
0.0 50.3 414
0.0 49.4 297
12
36
28
0.0 49.7 24
19.29
4 18
1992
0.017
808
19.29
4.18
1991
0.017
209
0.0
43.4
18 36
4.55
1990
0 017
414
0.0
43.0
21.03
5.27
1989
0.014
776
0.0
45.4
18 36
5.88
1986
0.015
213
20 38
708
1978
0.011
137
4^

Table A. 1. (cont'd)
Solids
Bulk
Total
Inorg
Organic
deposition
Total
Sampling
dry wt
Density
Carbon
Carbon
Carbon
Pb-210
Cs-137
period
rate
Mercury
Sample I D.
Date
g/g
g/cmA3
%
%
%
pCi/g
PCi/g
year e/cmA2-vr
ng/g
WCA1:37:12-14
07/15/92
0.069
0.067
46.9
6.69
4.45
1959
0.018
95
WCA1:37:14-16
07/15/92
0.063
0.062
46.7
5.61
4.46
1950
0.016
73
WCA1:37:16-18
07/15/92
0.072
0.072
47.4
0.0
47.4
3.90
3.34
1942
0.018
92
WCA1:37:18-20
07/15/92
0.068
0.069
47.5
2.22
2.19
1932
0.023
53
WCAl:37:20-22
07/15/92
0.072
0.070
2 58
1.83
1926
0.016
37
WCAl:37:22-24
07/15/92
0.078
0.077
2.13
1 49
1916
0.015
21
WCAl:37:24-26
07/15/92
0.067
0.068
49.0
1.46
0.92
1903
0.014
25
WCAl:37:26-28
07/15/92
0.078
0.076
0.43
1.10
1892
0.034
59
WCAl:37:28-30
07/15/92
0081
0.081
0 73
0.24
1887
0.017
33
WCA1:37:30-32
07/15/92
0.085
0,084
47.7
1.35
0 88
1876
0.007
42
WCAl:37:32-34
07/15/92
0.079
0080
0.37
0.63
1825
0.005
33
WCAl:37:34-36
07/15/92
0.093
0.089
28
WCA1:37:36-38
07/15/92
0086
0.083
50.9
31
WCA1:37:38-40
07/15/92
0.086
0.083
19
WCAl:37:40-42
07/15/92
NA
NA
NA
WCAl:37:42-44
07/15/92
0.083
0.080
51.1
20
WCAl:37:44-46
07/15/92
NA
NA
NA
WCAl:37:46-48
07/15/92
NA
NA
NA
WCAl:37:48-50
07/15/92
0.090
0.093
48.4
18
WCAl:37:50-52
07/15/92
0.094
0,091
38
WCA1:37:52-54
07/15/92
0.100
0.100
17
WCA1:37:54-56
07/15/92
0.105
0.099
50.0
0.0
49.9
13DL
WCA1:37:56-57
07/15/92
0.108
0.103
15
WCA1:38:00-02
07/15/92
0.093
0.070
9.23
1.14
1992
0.063
324
WCA1:38:02-04
07/15/92
0.087
0.085
43.1
13.84
111
1990
0.039
317
WCA1:38:04-06
07/15/92
0.095
0.095
16.30
1.15
1985
0.028
339
WCA1:38:06-08
07/15/92
0.065
0.065
19.69
0.95
1977
0.019
381
WCA1:38:08-10
07/15/92
0.111
0.112
12.29
6.40
1970
0023
198
WCA1:38:10-12
07/15/92
0.087
0.085
12.64
6.74
1958
0016
184
WCA1:38:12-14
07/15/92
0.086
0.086
10.19
5.16
1945
0.013
213
WCA1:38:14-16
07/15/92
0.082
0.083
6.04
2.56
1929
0.013
109
WCA1:38:16-18
07/15/92
0.077
0.077
5.57
1.63
1913
0.009
186
WCA1:38:18-20
07/15/92
0.078
0.078
2.69
1 90
1888
0008
97
WCA1:38:20-22
07/15/92
0.073
0.072
49.3
1.74
1.01
1860
0.005
86

Table A. 1. (cont'd)
Solids
Bulk
Total
Sampling
dry wt
Density
Carbon
Sample I D.
Date
g/g
g/cmA3
%
WCAl -.38:22-24
07/15/92
0.080
0.079
WCA1:38:24-26
07/15/92
0.071
0.069
WCAl :38:26-28
07/15/92
0.074
0.075
WCAl :38.28-30
07/15/92
0.078
0.075
WCAl :38:30-32
07/15/92
0.073
0.072
WCAl:38:32-34
07/15/92
0.072
0.071
WCAl:38:34-36
07/15/92
0.074
0.071
50.4
WCAl :38:36-38
07/15/92
0.081
0.078
WCAl:38:38^tO
07/15/92
0.081
0.079
WCAl:38:40-42
07/15/92
0.090
0.091
WCAl :38:42-44
07/15/92
0.104
0.104
WCAl :38:44-46
07/15/92
0.115
0.109
WCAl :38:46-47
07/15/92
0.113
0.110
WCAl :39:00-02
07/15/92
0.041
0.040
WCAl :39:02-04
07/15/92
0.086
0.085
WCA1:39:04-06
07/15/92
0084
0.085
WCAl:39:06-08
07/15/92
0.085
0.086
WCAl :39:08-10
07/15/92
0.084
0.085
WCAl :39:10-12
07/15/92
0.079
0.080
WCAl :39:12-14
07/15/92
0.078
0.079
WCAl :39:12-14
07/15/92
0.086
0.085
WCAl :39:14-16
07/15/92
0078
0.078
WCAl :39:16-18
07/15/92
0.079
0.079
WCAl :39:18-20
07/15/92
0.079
0.079
WCAl :39:20-22
07/15/92
0.079
0.077
WCAl :39:22-24
07/15/92
0.079
0.080
WCAl :39:24-26
07/15/92
0.073
0.076
WCAl :39:26-28
07/15/92
0.084
0.085
WCAl :39:28-30
07/15/92
0.083
0.084
WCAl :39:30-32
07/15/92
0089
0089
WCAl 39:32-34
07/15/92
0 098
0 099
WCAl:39:34-36
07/15/92
0.104
0.107
WCAl:39:36-37
07/15/92
0.100
0.100
WCAl :40:00-02
07/15/92
0.025
0.025
Inorg.
Carbon
128
65
63
66
67
48
27
43
44
97
78
90
526
357
231
268
129
150
161
138
100
127
87
104
126
104
89
81
69
62
74
83
1236 7.30 1992 0.029 413
Organic
Carbon
deposition
Pb-210 Cs-137 period rate
pCi/g pCi/g year g/cmA2-yr
0.34 0.84 1804 0.005
Total
Mercury
OgZg
95
147

Table A l. (cont'd)
Solids
Bulk
Total
Inorg.
Organic
deposition
Total
Sampling
dry wt
Density
Carbon
Carbon
Carbon
Pb-210
Cs-137
period
rate
Mercury
Samóle I D.
Date
g/g
ii/cmA3
%
%
%
pCi/g
pCi/g
year 2/cmA2-vr
ng/g
WCAl :40:02-04
07/15/92
0.052
0.050
51.4
11.60
4.92
1990
0.029
159
WCA1:40:04-06
07/15/92
0.104
0,104
9.93
5 35
1986
0030
134
WCAl :40:06-08
07/15/92
0.083
0.084
11.54
6.01
1979
0020
167
WCAl :40:08-10
07/15/92
0088
0.089
14.12
5.93
1969
0.013
119
WCAl: 40:10-12
07/15/92
0 098
0.098
5.76
4.22
1951
0.017
120
WCA1:40:12-14
07/15/92
0.096
0.098
3,79
2.89
1937
0.017
46
WCAl :40:14-16
07/15/92
0.077
0.079
3.23
1.64
1923
0.013
47
WCA1:40:16-18
07/15/92
0.071
0.071
2.59
1.63
1908
0.010
62
WCAl :40:18-20
07/15/92
0.079
0.078
50.9
1.95
NA
1889
0.008
54
WCAl:40:20-22
07/15/92
0.083
0.085
098
1.01
1855
0.005
31
WCAl:40:22-24
07/15/92
0.084
0088
000
19
WCAl:40:24-26
07/15/92
0076
0077
33
WCAl:40:26-28
07/15/92
0.076
0078
34
WCAl :40:28-30
07/15/92
0.068
0.070
38
WCAl:40:30-32
07/15/92
0.068
0.070
24
WCAl :40:32-34
07/15/92
0073
0.072
35
WCAl :40:34-36
07/15/92
0.068
0.067
37
WCAl :40:36-38
07/15/92
0.063
0064
504
26
WCAl :40:38-40
07/15/92
0.065
0.066
11
WCAl:40:40-42
07/15/92
0.071
0.074
23
WCAl .40:42-44
07/15/92
0085
0.087
19
WCAl :40:44-46
07/15/92
0.098
0 102
17
WCAl :41:00-02
07/15/92
0.013
0.013
179
WCAl :41:02-04
07/15/92
0.033
0,034
135
WCAl :41:04-06
07/15/92
0.102
0.103
178
WCAl :41:06-08
07/15/92
0.087
0.087
137
WCAl 41:08-10
07/15/92
0.066
0.067
197
WCAl :41:10-12
07/15/92
0.075
0077
137
WCA1:41:12-14
07/15/92
0.069
0.072
78
WCAl :41:14-16
07/15/92
0.065
0.066
67
WCAl :41 16-18
07/15/92
0 067
0.071
51
WCAl :41:18-20
07/15/92
0.074
0.076
47
WCAl :41:20-22
07/15/92
0.080
0083
180
WCAl :41:22-24
07/15/92
0.082
0084
77

Table A. 1. (cont'd)
Solids
Bulk
Total
Sampling
dry wt
Density
Carbon
Sample I D.
Date
&1&
g/cmA3
%
WCAl:41:24-26
07/15/92
0083
0.085
WCAl:41:26-28
07/15/92
0.077
0082
WCAl:41:28-30
07/15/92
0.094
0.097
WCAl:41:30-32
07/15/92
0.086
0 090
WCAl:41:32-34
07/15/92
0.089
0.091
WCAl:41:34-36
07/15/92
0.095
0.097
WCAl:42:00-02
07/15/92
0.088
0.091
WCA1:42:02-04
07/15/92
0.087
0.089
WCA1:42:04-06
07/15/92
0.085
0.089
WCA1:42:06-08
07/15/92
0.075
0.077
WCA1:42:08-10
07/15/92
0.072
0.074
WCA1:42:10-12
07/15/92
0.093
0.095
WCA1:42:12-14
07/15/92
0.084
0.087
WCA1:42:14-16
07/15/92
0.091
0095
WCA1:42:16-18
07/15/92
0.063
0.066
WCAl:42:18-20
07/15/92
0.057
0.059
WCAl:42:20-22
07/15/92
0.072
0,073
WCA1:42:22-24
07/15/92
0.070
0.071
WCA1:42:24-26
07/15/92
0065
0.068
WCAl:42:26-28
07/15/92
0.063
0.065
WCAl:42:28-30
07/15/92
0.076
0.079
WCAl:42:30-32
07/15/92
0.070
0.074
WCA1:42:32-34
07/15/92
0.064
0.066
WCAl:42:34-36
07/15/92
0.086
0.086
WCAl:42:36-38
07/15/92
0.091
0.095
WCAl:42:38-40
07/15/92
0.118
0.122
WCAl:42:40-42
07/15/92
0.101
0.102
Inorg. Organic
Carbon Carbon Pb-210
% % pCi/g
deposition
Cs-137 period rate
pCi/g year g/cmA2-yr
Total
Mercury
ng/g
30
31
16
17
26
5
137
276
171
138
113
68
64
70
51
39
32
61
51
20
58
47
51
27
25
28
23
â– ti
VO

Table A.2. Sediment and Mercury Analyses - Water Conservation Area 2
Solids
Bulk
Total
Inorg.
Organic
deposition
Total
Sampling
dry wt
Density
Carbon
Carbon
Carbon
Pb-210
Cs-137
period
rate
Mercury
Sample I D.
Date
S¿&
g/cmA3
%
%
%
pCi/g
pCi/g
year
u/cmA2-vr
ng/g
WCA2:25:00-02
03/11/92
0.017
0.017
12.13
0.44
1992
0.039
312
WCA2:25:02-04
03/11/92
0.010
0.010
38.0
13.76
0.39
1991
0.033
522
WCA2:25:04-06
03/11/92
0.014
0.014
11.06
0.46
1991
0.041
282
WCA2:25:06-08
03/11/92
0.049
0.049
17.22
1.45
1990
0.026
148
WCA2:25:08-10
03/11/92
0.106
0.109
20.29
5.30
1986
0019
129
WCA2:25:10-12
03/11/92
0.145
0 147
11.45
5.53
1972
0.022
98
WCA2:25:12-14
03/11/92
0.137
0.145
6.74
5,38
1954
0.022
86
WCA2:25:14-16
03/11/92
0.121
0.128
2.79
5.44
1937
0.030
69
WCA2:25:16-18
03/11/92
0.134
0.143
3.11
4.32
1927
0.020
64
WCA2:25:18-20
03/11/92
0.134
0.135
3.19
3 29
1908
0.011
62
WCA2:25:20-22
03/11/92
0.142
0 145
47.4
0.00
2.95
68
WCA2:25:22-24
03/11/92
0.164
0.180
0.73
1.37
1862
0.011
31
WCA2:25:24-26
03/11/92
0.116
0 121
0.00
16
WCA2:25:26-28
03/11/92
0.100
0 105
30
WCA2:25:28-30
03/11/92
0.109
0.109
17
WCA2:25:30-32
03/11/92
0.171
0.174
24
WCA2:25:32-34
03/11/92
0.115
0.118
47.9
45
WCA2:25:34-36
03/11/92
0.126
0.133
40
WCA2:26A:00-02
03/11/92
0.048
0.050
34.1
3.4
30.8
11.72
1.57
1992
0.029
157
WCA2:26A:02-04
03/11/92
0.237
0.249
37.1
9.23
1.97
1988
0.033
78
WCA2:26A:04-06
03/11/92
0.146
0.146
43.5
6.09
1.38
1968
0.026
67
WCA2:26A:06-08
03/11/92
0.112
0.116
46.0
0.1
46.0
1.32
0.50
1954
0.079
47
WCA2:26A:08-10
03/11/92
0.109
0.116
47.0
0.0
47.0
1.29
0.34
1951
0.073
48
WCA2:26A:10-12
03/11/92
0.141
0.144
46.3
1.82
0.01
1948
0.047
95
WCA2:26A:12-14
03/11/92
0.183
0.189
47.4
2.25
0.36
1941
0.031
29
WCA2:26A:14-16
03/11/92
0.179
0.190
40.9
1.21
0.26
1925
0.035
17
WCA2:26A:16-18
03/11/92
0.151
0 166
35.0
0.0
35.0
1.04
BDL
1912
0.027
13
WCA2:26A: 18-20
03/11/92
0.168
0 169
47.7
0.0
47.7
0.87
BDL
1896
0.020
25
WCA2:26A:20-22
03/11/92
0.175
0 185
0.70
0.09
1872
0.012
31
WCA2:26A:22-24
03/11/92
0.172
0.206
41.5
0.0
41.5
30
WCA2:26B:00-02
03/11/92
0.013
0014
227
WCA2:26B:02-04
03/11/92
0.045
0047
214
WCA2:26B:04-06
03/11/92
0.237
0.259
79
WCA2:26B:06-08
03/11/92
0.115
0 112
46

Table A.2. (cont'd)
Solids
Bulk
Total
Inorg.
Organic
deposition
Total
Sampling
dry wt
Density
Carbon
Carbon
Carbon
Pb-210
Cs-137
period rate
Mercury
Sample I D.
Date
e¿&
e/cmA3
%
%
%
pCi/R
pCj/g
year e/cmA2-vr
0g/g
WCA2:26B:08-10
03/11/92
0.127
0.132
41
WCA2:26B:10-12
03/11/92
0.136
0 143
62
WCA2:26B:12-14
03/11/92
0.147
0 157
36
WCA2:26B:14-16
03/11/92
0.170
0 170
18
WCA2:26B:16-18
03/11/92
0.125
0.129
31
WCA2:26B: 18-20
03/11/92
0.259
0.295
12
WCA226B20-22
03/11/92
0.170
0 182
37
WCA2:27:02-04
03/12/92
0.174
0 169
158
WCA2:27:04-06
03/12/92
0.157
0.167
117
WCA2:27:06-08
03/12/92
0.121
0 125
144
WCA2:27:08-10
03/12/92
0.135
0.143
80
WCA2:27:10-12
03/12/92
0.124
0.125
79
WCA2:27:12-14
03/12/92
0.142
0 149
68
WCA2:27:14-16
03/12/92
0.158
0 159
47
WCA2:27:16-18
03/12/92
0.123
0.127
60
WCA2:27:18-20
03/12/92
0.117
0.115
64
WCA2:27:20-22
03/12/92
0.117
0.119
35
WCA2:27:22-24
03/12/92
0.126
0.124
33
WCA2:27:24-26
03/12/92
0.109
0 109
48
WCA2:27:26-28
03/12/92
0.133
0.140
31
WCA2:27:28-30
03/12/92
0.118
0.117
35
WCA2:27:30-32
03/12/92
0.148
0.147
36
WCA2:27:32-34
03/12/92
0.148
0 143
35
WCA2:27:34-36
03/12/92
0.131
0.130
49
WCA2:28:00-02
03/12/92
0.139
0.140
77
WCA2:28:02-04
03/12/92
0.118
0.120
63
WCA2:2 8:04-06
03/12/92
0.105
0.106
NA
WCA2:28:06-08
03/12/92
0.090
0.091
45
WCA2:28;08-10
03/12/92
0,080
0.082
51
WCA2:28:10-12
03/12/92
0095
0 098
44
WCA2:28:12-14
03/12/92
0.104
0 107
NA
WCA2:28:14-16
03/12/92
0 108
0 112
28
WCA2:28:16-18
03/12/92
0 101
0 099
52
WCA2:28:18-20
03/12/92
0.098
0 103
31

Table A.2. (cont'd)
Sample I D
WCA2:28:20-22
WCA2:28:22-24
WCA2:28:24-26
WCA2:28:26-28
WCA2:28:28-31
Sampling
Date
03/12/92
03/12/92
03/12/92
03/12/92
03/12/92
Solids
dry wt
g/g
0.104
0.115
0.113
0.117
0.122
Bulk
Density
g/cmA3
0.106
0.119
0.112
0.118
0.118
Total
Carbon
%
Inorg.
Carbon
%
WCA2:29:00-02
03/12/92
0.072
0.077
44.6
0.8
WCA2:29:02-04
03/12/92
0.094
0.094
45.5
WCA2:29:04-06
03/12/92
0.067
0.070
46.3
0.7
WCA2:29:06-08
03/12/92
0.122
0.135
45.4
WCA2:29:08-10
03/12/92
0.108
0.111
47 8
WCA2:29:10-12
03/12/92
0.065
0070
47.8
0.1
WCA2:29:12-14
03/12/92
0.049
0.043
483
0.0
WCA2:29:14-16
03/12/92
0.105
0.110
47.7
WCA2:29:16-18
03/12/92
0.082
0.090
45.0
WCA2:29:18-20
03/12/92
0.073
0.089
46.6
WCA2:29:20-22
03/12/92
0.075
0.072
NA
WCA2:29:22-24
03/12/92
0.198
0.199
NA
WCA2:29:24-26
03/12/92
0.146
0.152
45.6
0.0
WCA2:29:26-28
03/12/92
0.213
0.202
NA
WCA2:29:28-30
03/12/92
0.138
0.131
NA
WCA2:29:30-32
03/12/92
0 158
0 155
483
WCA2:29:32-34
03/12/92
0 101
0.103
NA
WCA2:29:34-36
03/12/92
0.176
0 187
NA
WCA2:29:36-38
03/12/92
0.140
0.138
50.3
0.0
WCA2:29:38-40
03/12/92
0.112
0.112
NA
WCA2:29:40-42
03/12/92
0.150
0.150
44.5
0.0
WCA2:30:00-02
03/12/92
0 138
0.143
45.3
0.4
WCA2:30:02-04
03/12/92
0.026
0.023
46.6
0.2
WCA2:30:04-06
03/12/92
0.046
0.045
46.7
0.4
WCA2:30:06-08
03/12/92
0.026
0.025
44.1
0.1
WCA2:30:08-10
03/12/92
0.084
0.089
48.2
WCA2:30:10-12
03/12/92
0.098
0.097
45.7
WCA2:30:12-14
03/12/92
0.098
0.098
47.9
0,0
WCA2:30:14-16
03/12/92
0 127
0 142
46.7
0.0
Organic deposition Total
Carbon Pb-210 Cs-137 period rate Mercury
% pCi/g pCi/g year g/cmA2-yr ng/g
103
36
37
45
52
43.8
4.08
0.74
1992
0.080
74
4.08
0.74
1990
0.075
56
45.7
5.47
0.83
1987
0.052
79
5.47
0.83
1985
0,047
83
4.73
0.29
1978
0.045
59
47.7
4.73
0.29
1973
0.038
145
48.3
4.29
0.90
1969
0.037
109
4.29
0.90
1967
0.035
42
3.40
1.66
1960
0.035
33
3.40
1.66
1954
0.029
156
2.64
7.85
1947
0.031
101
2.64
7.85
1942
0.026
49
45.6
0.85
2.99
1922
0.043
48
0.53
1.74
1914
0.054
NA
0.67
1.74
1905
0.033
45
060
0.32
1896
0.027
43
0.53
0.01
1882
0 020
20
0.53
0.01
1870
0014
22
50.3
0.12
0.28
15
BDI.
30
44.6
UD1.
44.9
66
46.4
358
46.4
172
44.0
313
120
82
47.8
95
46.7
55
K)

Table A.2. (cont'd)
Solids
Bulk
Total
Inorg
Organic
deposition
Total
Sampling
dry wt
Density
Carbon
Carbon
Carbon
Pb-210
Cs-137
period rate
Mercury
Sample I D.
Date
g/g
e/cmA3
%
%
%
pCj/g
P-Cj/g
year e/cmA2-vr
n a/g.
WCA2:30:16-18
03/12/92
0.090
0.095
48.1
0.0
48.1
75
WCA2:30:18-20
03/12/92
0.133
0.133
49 1
59
WCA2:30:20-22
03/12/92
0.136
0 134
49 5
50
WCA2:30:22-24
03/12/92
0.064
0067
34
WCA2:30:24-26
03/12/92
0.122
0.120
17
WCA2:30:26-28
03/12/92
0.132
0 123
26
WCA2:30:28-30
03/12/92
0.119
0.126
50.8
27
WCA2:30:30-32
03/12/92
0.122
0.129
27
WCA2:30:32-34
03/12/92
0.111
0.116
60
WCA2:30:34-36
03/12/92
0 167
0.169
52 4
oo
52.4
20
BDL, below detection limit
NA, analysis not available

Table A.3. Sediment and Mercury Analyses - Water Conservation Area 3
Solids
Bulk
Total
Inorg.
Organic
deposition
Total
Sampling
dry wt
Density
Carbon
Carbon
Carbon
Pb-210
Cs-137
period
rate
Mercury
Sample I D
Date
e/cmA3
%
%
%
pCi/g
pCi/g
year
e/cmA2-vr
DSl&
WCA3:01:00-02
01/21/92
0.035
0.036
101
WCA3:01:02-04
01/21/92
0.077
0.074
79
WCA3:01:04-06
01/21/92
0.111
0 113
109
WCA3:01:06-08
01/21/92
0.111
0.100
111
WCA3:01:08-10
01/21/92
0.108
0.111
89
WCA3:01:10-12
01/21/92
0 101
0.096
85
WCA3.01 12-14
01/21/92
0 106
0 092
66
WCA3:01:14-16
01/21/92
0 110
0 097
49
WCA3:01:16-18
01/21/92
0.107
0.086
26
WCA3:01:18-20
01/21/92
0.095
0.082
38
WCA3:01:20-22
01/21/92
0.109
0.108
40
WCA3:01:22-24
01/21/92
0.143
0.150
31
WCA3:01:24-26
01/21/92
0.157
0.161
44
WCA3:01:26-28
01/21/92
0.137
0 149
64
WCA3:01:28-30
01/21/92
0 112
0.154
32
WCA3:02:00-02
01/21/92
0.035
0.028
101
WCA3:02.02-04
01/21/92
0.084
0.073
81
WCA3:02:04-06
01/21/92
0 101
0.107
61
WCA3:02:06-08
01/21/92
0.109
0.091
47
WCA3:02:08-10
01/21/92
0.096
0.082
55
WCA3:02:10-12
01/21/92
0.095
0083
29
WCA3:02:12-14
01/21/92
0088
0.080
30
WCA3:02:14-16
01/21/92
0.097
0.085
19
WCA3:02:16-18
01/21/92
0.097
0.086
19
WCA3:02:18-20
01/21/92
0.105
0.101
18
WCA3:02:20-22
01/21/92
0.115
0.117
16
WCA3:02:22-24
01/21/92
0.154
0 142
23
WCA3:02:24-26
01/21/92
0.141
0.126
32
WCA3:02:26-28
01/21/92
0.167
0.167
43
WCA3:02:28-30
01/21/92
0.148
0.147
45
WCA3:02:30-32
01/21/92
0.106
0.114
17
WCA3:02:32-34
01/21/92
0.102
0.107
18
WCA3:02:34-35
01/21/92
0 171
0 120
25
WCA3:13:00-02
03/10/92
0.173
0.175
15.22
2 16
1992
0.035
81

Table A.3. (cont'd)
Solids
Bulk
Total
Inorg.
Organic
deposition
Total
Sampling
dry wt
Density
Carbon
Carbon
Carbon
Pb-210
Cs-137
period
rate
Mercury
Sample I D
Date
ate
g/cmA3
%
%
%
pCj/g
pCi/R
year
g/cmA2-vr
na/g.
WCA3:13:02-04
03/10/92
0.176
0.186
32.3
11.22
3.88
1980
0.033
74
WCA3:13:04-06
03/10/92
0.153
0.159
11.41
5 03
1966
0 021
86
WCA3:13:06-08
03/10/92
0.125
0.126
4.14
1.22
1946
0.031
77
WCA3:13:08-10
03/10/92
0.137
0 136
2 93
1.33
1936
0.032
55
WCA3:13:10-12
03/10/92
0.125
0.128
1 98
1.24
1926
0.035
24
WCA3:13:12-14
03/10/92
0.124
0 126
095
1.24
1918
0.056
24
WCA3:13:14-16
03/10/92
0 143
0.144
1.27
020
1913
0.036
29
WCA3:13:16-18
03/10/92
0.173
0.179
31.8
1.53
0.13
1904
0022
31
WCA3:13:18-20
03/10/92
0.245
0.263
1.06
0.01
1882
0.016
30
WCA3:13:20-22
03/10/92
0.199
0.208
27.7
32
WCA3:13:22-24
03/10/92
0.220
0.239
39
WCA3:14:00-02
03/10/92
0.036
0.033
44.2
0.0
44.1
347
WCA3:14:02-04
03/10/92
0.087
0.088
43.7
231
WCA3:14:04-O6
03/10/92
0.118
0.125
44.2
0.0
44 2
123
WCA3:14:06-08
03/10/92
0.044
0.046
48.5
0.0
48.5
291
WCA3:14:08-10
03/10/92
0.095
0.097
49.0
0.0
49.0
no
WCA3:14:10-12
03/10/92
0.117
0.106
50.8
51
WCA3:14:12-14
03/10/92
0.114
0.115
45.5
0.0
45.5
172
WCA3:14:14-16
03/10/92
0.130
0,124
46.6
67
WCA3:14:16-18
03/10/92
0.145
0.147
47.3
0.0
47.3
95
WCA3:14:18-20
03/10/92
0.105
0.104
52.7
83
WCA3:14:20-22
03/10/92
0099
0 097
61
WCA3:14:22-24
03/10/92
0204
0.216
16
WCA3:14:24-26
03/10/92
0.236
0.266
25.4
BDL
WCA3:14:26-28
03/10/92
0.297
0.346
BDL
WCA3:14:28-30
03/10/92
0.358
0.462
17.3
9.5
7.8
BDL
WCA3:14:30-32
03/10/92
0.298
0359
BDL
WCA3:14:32-34
03/10/92
0.369
0.440
16.8
102
6.6
BDL
WCA3:15:00-02
03/10/92
0.035
0.030
20.90
1.73
1992
0.032
66
WCA3:15:02-04
03/10/92
0.093
0084
396
21.89
1 88
1990
0029
87
WCA3:15:04-06
03/10/92
0 103
0 095
400
2026
2 95
1984
0.026
221
WCA3:1S:06-08
03/10/92
0 150
0.136
21.01
5 73
1975
0019
137
WCA3:15:08-10
03/10/92
0 176
0.164
10.22
564
1956
0.022
81
WCA3:15:10-12
03/10/92
0 193
0 181
51.7
4 70
2.77
1936
0.025
57

Table A. 3. (cont'd)
Solids
Bulk
Total
Inorg
Organic
deposition
Total
Sampling
dry wt
Density
Carbon
Carbon
Carbon
Pb-210
Cs-137
period
rate
Mercury
Sample I D.
Date
g/g
e/cmA3
%
%
%
pCi/g
pCi/g
year
a/cmA2-yr
Ü&1&
WCA3:15:12-14
03/10/92
0.174
0.160
3.20
1.90
1916
0.020
95
WCA3:15:14-16
03/10/92
0.153
0.140
3 06
0 85
1893
0.010
40
WCA3;15:16-18
03/10/92
0.165
0.152
0.28
045
1830
0.015
BDL
WCA3:15:18-20
03/10/92
0.132
0.120
0.22
0.007
18
WCA3:15:20-22
03/10/92
0.162
0.150
14
WCA3:15:22-24
03/10/92
0.140
0.127
BDL
WCA3:15:24-26
03/10/92
0.150
0,136
34
WCA3:16:00-02
03/10/92
0.018
0.018
239
WCA3:16:02-04
03/10/92
0.173
0.181
32.1
99
WCA3:16:04-06
03/10/92
0.056
0.057
306
WCA3:16:06-08
03/10/92
0.188
0.198
92
WCA3:16:08-10
03/10/92
0.168
0.170
57
WCA3:16:10-12
03/10/92
0.217
0.228
65
WCA3:16:12-14
03/10/92
0.237
0.244
49
WCA3:16:14-16
03/10/92
0210
0.212
48.9
87
WCA3:16:16-18
03/10/92
0.215
0217
16
WCA3:16:18-20
03/10/92
0.197
0.202
12
WCA3:16:20-22
03/10/92
0.166
0.172
52.4
24
WCA3:16:22-24
03/10/92
0.190
0.195
33
WCA3:17:00-02
03/10/92
0.082
0.082
44.7
0.0
44.7
402
WCA3:17:02-04
03/10/92
0.158
0.172
41.7
294
WCA3:17:04-06
03/10/92
0.120
0,128
46.3
0.0
46.3
303
WCA3:17:06-08
03/10/92
0.165
0.166
44.2
128
WCA3:17:08-10
03/10/92
0.099
0.105
209
WCA3:17:10-12
03/10/92
0.158
0.168
46.5
0.0
46.5
122
WCA3:17:12-14
03/10/92
0.135
0.137
45.3
113
WCA3:17:14-16
03/10/92
0.122
0.125
50.0
106
WCA3:17:16-18
03/10/92
0.145
0 150
45.9
0.0
45.9
59
WCA3:18:00-02
03/11/92
0.015
0.014
38.3
0.0
38.3
566
WCA3:18:02-04
03/11/92
0.118
0.117
42.4
0.0
42.4
131
WCA3:18:04-06
03/11/92
0.157
0.169
42.2
0.0
42.2
304
WCA3:18:0608
03/11/92
0.224
0.233
44.3
0.0
44.3
228
WCA3:18:08-10
03/11/92
0.170
0.183
49.8
0.0
49.8
225
WCA3:18:10-12
03/11/92
0.172
0 178
45.8
121

Table A. 3. (cont'd)
Solids
Bulk
Total
Inorg.
Sampling
dry wt
Density
Carbon
Carbon
Sample I D.
Date
&Zg
g/cmA3
%
%
WCA3:18:12-14
03/11/92
0.131
0.130
49 9
WCA3:18:14-16
03/11/92
0.135
0 143
480
WCA3:18:16-18
03/11/92
0.150
0.135
49.1
0.0
WCA3:18:18-20
03/11/92
0.139
0.137
49.0
0.0
WCA3:18:20-22
03/11/92
0.112
0.114
WCA3:18:22-24
03/11/92
0.089
0.092
WCA3:18:24-26
03/11/92
0.105
0.105
48.7
0.0
WCA3:18:26-28
03/11/92
0.107
0.106
WCA3:18:28-30
03/11/92
0.098
0 098
WCA3:18:30-32
03/11/92
0.097
0.099
45.9
0.0
WCA3:18:32-34
03/11/92
0.125
0 125
WCA3:18:34-36
03/11/92
0.123
0.118
WCA3:18:36-38
03/11/92
0.107
0.102
WCA3:18:3840
03/11/92
0.108
0 106
47.8
WCA3:18:40-42
03/11/92
0.113
0.113
WCA3:18:4244
03/11/92
0.101
0.098
47.0
0.0
WCA3:18:4446
03/11/92
0.125
0 122
WCA3:18:4648
03/11/92
0.097
0.094
WCA3:18:48-50
03/11/92
0 120
0.119
WCA3:18:50-52
03/11/92
0.139
0.140
WCA3:18:52-53
03/11/92
0.144
0.144
WCA3:19:00-02
03/11/92
0.043
0.043
WCA3:19:02-04
03/11/92
0.165
0.171
WCA3:19:04-06
03/11/92
0.176
0.191
WCA3:19:06-08
03/11/92
0.167
0.176
WCA3:19:08-10
03/11/92
0.166
0.176
WCA3:19:10-12
03/11/92
0.117
0.119
WCA3:19:12-14
03/11/92
0.112
0.118
WCA3:19:14-16
03/11/92
0.097
0 094
WCA3:19:16-18
03/11/92
0.110
0.114
WCA3:19:18-20
03/11/92
0.111
0.115
WCA3:19:20-22
03/11/92
0.128
0.129
WCA3:19:22-24
03/11/92
0.104
0 105
WCA3:19:24-26
03/11/92
0.106
0.109
Organic deposition Total
Carbon Pb-210 Cs-137 period rate Mercury
%
49 1
490
48.7
45 9
47.0
pCi/g pCi/g year g/cmA2-yr ng/g
84
80
42
69
57
45
62
49
145
121
103
90
115
80
85
52
70
18
52
61
52
1.54
0.60
1992
0.259
40
3.52
0.93
1992
0.112
25
6.32
3.29
1988
0.056
57
10.36
5.90
1981
0.027
194
10.54
7.78
1964
0.016
152
5 83
7.17
1927
0.009
142
1.06
5.74
1872
0.009
105
030
2 83
1818
0.006
92
0.00
1.18
62
36
32
30
28

Table A.3. (cont'd)
Solids
Bulk
Total
Inorg.
Organic
deposition
Total
Sampling
dry wt
Density
Carbon
Carbon
Carbon
Pb-210
Cs-137
period rate
Mercury
Sample I D
Date
g/g
g/cmA3
%
%
%
pCi/g
pCi/g
year g/cmA2-vr
ng/a
WCA3:19:26-28
03/11/92
0.140
0.137
28
WCA3:19:28-30
03/11/92
0 184
0 190
38
WCA3:19:30-32
03/11/92
0.162
0 165
45
WCA3:20A: 02-04
03/11/92
0.379
0.447
13
WCA3:20B:02-04
03/11/92
0.601
0.943
BDL
WCA3:20C:02-04
03/11/92
0.622
0.648
20
WCA3:20C:06-08
03/11/92
0.635
0.874
BDL
WCA3:20D:02-04
03/11/92
0.756
1.149
BDL
WCA3:20D:06-08
03/11/92
0.638
0.740
10
WCA3:21:00-02
03/11/92
0.169
0.175
72
WCA3:21:02-04
03/11/92
0.239
0.251
58
WCA3:21:04-06
03/11/92
0.270
0.302
62
WCA3:21:06-08
03/11/92
0.386
0 468
33
WCA3:21:08-10
03/11/92
0.537
0.735
18
WCA3:21:10-12
03/11/92
0485
0.648
BDL
WCA3:22:00-02
03/11/92
0.030
0.028
27
WCA3:22:02-04
03/11/92
0 116
0.110
87
WCA3:22:04-06
03/11/92
0.182
0.193
124
WCA3:22:06-08
03/11/92
0.178
0.183
74
WCA3:22:08-10
03/11/92
0.156
0.158
107
WCA3:22:10-12
03/11/92
0.189
0.201
47
WCA3:22:12-14
03/11/92
0.318
0380
17
WCA3:22:14-16
03/11/92
0.291
0.335
27
WCA3:22:16-18
03/11/92
0.249
0265
27
WCA3:22:18-20
03/11/92
0.261
0.289
50
WCA3:22:20-22
03/11/92
0.341
0.408
22
WCA3:23A:02-04
03/11/92
0.427
0.533
89
WCA3:23B:02-04
03/11/92
0.425
0.529
12
WCA3:23C:02-04
03/11/92
0.278
0.306
13
WCA3:23D 06-08
03/11/92
0 352
0 411
14
WCA3:23E:02-04
03/11/92
0.369
0.447
16
WCA3:24:00-02
03/11/92
0.018
0.017
331
WCA3:24:02-04
03/11/92
0.163
0 160 42.0
92
WCA3:24:04-06
03/11/92
0.118
0.120
117

Table A.3. (cont'd)
Solids
Bulk
Total
Inorg.
Organic
deposition
Total
Sampling
dry wt
Density
Carbon
Carbon
Carbon
Pb-210
Cs-137
period rate
Mercury
Sample I D.
Date
g/g
g/cmA3
%
%
%
pCi/g
pCi/g
vear g/cmA2-vr
ng^g
WCA3:24:06-08
03/11/92
0.109
0.105
66
WCA3:24:08-10
03/11/92
0.124
0.127
58
WCA3:24:10-12
03/11/92
0.126
0.132
83
WCA3:24:12-14
03/11/92
0 128
0 128
74
WCA3:24:14-16
03/11/92
0.141
0.144
89
WCA3:24:16-18
03/11/92
0,146
0.146
47.0
103
WCA3:24:18-20
03/11/92
0.112
0.113
85
WCA3:24:20-22
03/11/92
0.114
0.113
103
WCA3:24:22-24
03/11/92
0.113
0.111
65
WCA3:24:24-26
03/11/92
0.119
0.118
34
WCA3:24:26-28
03/11/92
0,116
0.112
26
WCA3:24:28-30
03/11/92
0.125
0.125
49.8
42
WCA3:24:30-32
03/11/92
0.131
0.129
47
WCA3:32:00-02
03/12/92
0.162
2.453
BDL
WCA3:32:02-04
03/12/92
0.187
2.696
12
WCA3:32:04-06
03/12/92
0.297
2.857
19
WCA3:32:06-08
03/12/92
0.271
2.552
34
WCA3:32:08-10
03/12/92
0.315
2.895
18
WCA3:32:10-12
03/12/92
0.320
2.900
42
WCA3:32:12-14
03/12/92
0.298
2.715
27
WCA3:32:14-16
03/12/92
0.282
2.530
28
WCA3:32:16-18
03/12/92
0.287
2 553
28
WCA3:32:18-20
03/12/92
0.276
2.443
21
WCA3:33:00-02
03/12/92
0.195
1.685
BDL
WCA3:33:02-04
03/12/92
0.275
2.575
BDL
WCA3:33:04-06
03/12/92
0.282
2 647
BDL
WCA3:33:06-08
03/12/92
0.314
3,081
15
WCA3:33:08-10
03/12/92
0.335
3.256
27
WCA3:33:10-12
03/12/92
0.315
2.843
33
WCA3:33:12-14
03/12/92
0.265
2.323
39
WCA3:33:14-16
03/12/92
0.256
2.231
63
WCA3:33:16-18
03/12/92
0.257
2.256
45
WCA3:33:18-20
03/12/92
0.247
2.228
47
WCA3:33:20-22
03/12/92
0.242
2 155
BDL

Table A.3. (cont'd)
Solids
Bulk
Total
Inorg.
Organic
deposition
Total
Sampling
dry wt
Density
Carbon
Carbon
Carbon
Pb-210
Cs-137
period rate
Mercury
Sample I D.
Date
a/cmA3
%
%
%
pCi/g
pCi/g
year g/cmA2-vr
mis
WCA3:33:22-24
03/12/92
0.232
2.075
BDL
WCA3:33:24-26
03/12/92
0.232
2.035
24
WCA3:33:26-28
03/12/92
0234
2053
29
WCA3:33:28-31
03/12/92
0.219
1.931
21
WCA3:C123:00-02
01/21/92
0.035
0.035
162
WCA3:C123:02-04
01/21/92
0.040
0.041
85
WCA3:C123:04-06
01/21/92
0 045
0 044
99
WCA3:C123:06-08
01/21/92
0.078
0.077
131
WCA3:C123:08-10
01/21/92
0.108
0.105
93
WCA3:C123:10-12
01/21/92
0.118
0.110
87
WCA3:C123.12-14
01/21/92
0.105
0.116
64
WCA3:C123:14-16
01/21/92
0.145
0.155
15
WCA3:C123:16-18
01/21/92
0.123
0.130
73
WCA3:C123:18-20
01/21/92
0.166
0.172
83
WCA3:C123:20-22
01/21/92
0.168
0.166
82
WCA3:C 123:22-24
01/21/92
0.181
0 182
81
WCA3:C123:24-26
01/21/92
0.177
0.191
77
WCA3:C 123:26-28
01/21/92
0 184
0219
80
WCA3:C123:28-30
01/21/92
0.372
0.413
37
WCA3:C123:30-34
01/21/92
0.382
0.973
42
WCA3:C123:34-38
01/21/92
0.399
1.000
26
WCA3:C123:3842
01/21/92
0.376
0.845
22
WCA3:C123:42-46
01/21/92
0.407
1.118
20
WCA3:C123:46-50
01/21/92
0.430
1.070
23
WCA3:C123:50-54
01/21/92
0.401
0.941
23
WCA3:C123:54-58
01/21/92
0.393
0.969
28
WCA3:C123:58-62
01/21/92
0.410
0.982
22
WCA3:C123:62-66
01/21/92
0.412
0.894
16
WCA3:C123:66-70
01/21/92
0.366
0,981
21
WCA3:C 123:70-74
01/21/92
0.374
1.050
21
BDL, below detection limit
NA, analysis not available

Table A.4. Sediment and Mercury Analyses - Everglades National Park
Solids
Bulk
Total
Inorganic
Organic
deposition
Total
Sampling
dry wt
Density
Carbon
Carbon
Carbon
Pb-210
Cs-137
period rate
Mercury
Sample I D.
ENP:01A:00-07
ENP:01A: 07-14
Date
03/09/92
03/09/92
ei&
0.277
0.440
a/cmA3
0.401
0689
%
%
%
pCi/g
pCi/g
vear e/cmA2-vr
14.6
37.8
ENP:01B:00-09
03/09/92
0.298
0.449
24.3
ENP:01B:09-15
03/09/92
0.410
0.630
25.1
ENP:02:00-02
03/09/92
0.198
0.200
110
ENP:02:02-04
03/09/92
0.174
0.186
43.7
85
ENP:02:04-06
03/09/92
0.145
0.154
101
ENP:02:06-08
03/09/92
0.131
0 126
105
ENP:02:08-10
03/09/92
0.150
0 146
65
ENP:02:10-12
03/09/92
0.130
0 129
46.8
82
ENP:02:12-14
03/09/92
0.115
0.112
75
ENP:02:14-16
03/09/92
0.124
0.126
61
ENP:02:16-18
03/09/92
0.137
0.137
99
ENP: 02:18-20
03/09/92
0.125
0 124
94
ENP:02:20-22
03/09/92
0.140
0.138
61
ENP:02:22-24
03/09/92
0.266
0.316
17
ENP:02:24-26
03/09/92
0.329
0.349
11
ENP:02:26-28
03/09/92
0.380
0.465
BDL
ENP:02:28-30
03/09/92
0.324
0.367
11
ENP:02:30-32
03/09/92
0.151
0.146
37
ENP:02:32-34
03/09/92
0.146
0.141
38
ENP:02:34-36
03/09/92
0.136
0.132
49
ENP:02:36-38
03/09/92
0.137
0.136
48
ENP:02:3840
03/09/92
0.128
0.121
28
ENP:02:40-42
03/09/92
0.164
0 165
22
ENP:02:4244
03/09/92
0.255
0.277
10
ENP:02:44-t6
03/09/92
0.296
0.318
BDL
ENP:02:46^»8
03/09/92
0.312
0 348
BDL
ENP:02:48-50
03/09/92
0.188
0.192
42.9
30
ENP:02:50-52
03/09/92
0.155
0.154
50
ENP:03A:00-03
03/09/92
0.662
0.647
41
ENP:03B:00-03
03/09/92
0.615
0.769
43
ENP: 04A: 00-08
03/09/92
0.339
0.391
43
ENP:04A:08-16
03/09/92
0.422
0482
38

Table A.4. (cont'd)
Solids
Bulk
Total
Inorganic
Organic
deposition
Total
Sampling
dry wt
Density
Carbon
Carbon
Carbon
Pb-210
Cs-137
period rate
Mercury
Sample I D
Date
g/g
e/cmA3
%
%
%
pCi/g
pCi/g
year g/cmA2-yr
Dg/g
ENP:04B:00-08
03/09/92
0.351
0.368
32
ENP:04B:08-16
03/09/92
0.472
0.527
19
ENP:04B: 16-24
03/09/92
0.511
0.610
NA
ENP:05:00-02
03/09/92
0.181
0.182
87
ENP:05:02-04
03/09/92
0.240
0.241
82
ENP:05:04-06
03/09/92
0.226
0229
92
ENP:05:06-08
03/09/92
0.253
0.265
82
ENP:05:08-10
03/09/92
0.256
0.274
103
ENP:05:10-12
03/09/92
0.259
0.276
93
ENP:05:12-14
03/09/92
0.277
0.307
82
ENP:05:14-16
03/09/92
0.271
0.304
81
ENP:05:16-18
03/09/92
0.271
0.304
85
ENP:05:18-20
03/09/92
0.279
0.314
72
ENP:05:20-22
03/09/92
0.324
0.385
65
ENP:05:22-25
03/09/92
0.337
0.401
59
ENP:06:00-02
03/09/92
0.279
0.324
17
ENP:06:02-04
03/09/92
0.298
0.334 15.2
19
ENP:06:04-06
03/09/92
0.346
0.418
25
ENP:06:06-08
03/09/92
0.328
0.394
20
ENP:06:08-10
03/09/92
0.367
0.447
NA
ENP:06:10-12
03/09/92
0.348
0.418
NA
ENP:06:12-14
03/09/92
0.320
0.389
17
ENP:06:14-16
03/09/92
0.338
0.389
16
ENP:06:16-18
03/09/92
0.285
0.345 15.7
15
ENP:06:18-20
03/09/92
0.308
0.378
14
ENP:06:20-22
03/09/92
0.312
0.387
11
ENP:06:22-24
03/09/92
0.343
0.376
19
ENP:06:24-26
03/09/92
0.158
0.160
50
ENP:06:26-28
03/09/92
0.153
0.146
36
ENP:06:28-30
03/09/92
0.193
0.186
38
ENP:06:30-32
03/09/92
0 193
0.206
28
ENP:06:32-34
03/09/92
0,254
0.273
10
ENP:06:34-36
03/09/92
0289
0,273
15
ENP:06:36-38
03/09/92
0 153
0.176
24

Table A.4.
(cont’d)
Solids
Bulk
Total
Inorganic
Sampling
dry wt
Density
Carbon
Carbon
Sample I D.
Date
g/g
g/cmA3
%
%
ENP:06:3840
03/09/92
0.321
0.403
ENP:06:40-42
03/09/92
0.410
0.518
ENP:06:42^t4
03/09/92
0.473
0.581
ENP:06:44-46
03/09/92
0.408
0.521
ENP:06:46^18
03/09/92
0.452
0.589
ENP:06:48-50
03/09/92
0.485
0.637
ENP:06:50-52
03/09/92
0.473
0.652
ENP:06:52-54
03/09/92
0.525
0.736
ENP:06:54-56
03/09/92
0.559
0.779
ENP:06:56-58
03/09/92
0.548
0.761
ENP:06:58-60
03/09/92
0.530
0.740
13.7
ENP:06:60-62
03/09/92
0.551
0.761
ENP:07:00-02
03/09/92
0.046
0.143
24.2
6.0
ENP:07:02-04
03/09/92
0.179
0,292
21.5
ENP:07:04-06
03/09/92
0.217
0.345
19.7
8 8
ENP:07:06-08
03/09/92
0.198
0.318
22.6
8.2
ENP:07:08-10
03/09/92
0.267
0.398
25.0
7.2
ENP:07:10-12
03/09/92
0.250
0370
22 8
7.9
ENP:07:12-14
03/09/92
NA
NA
284
5 9
ENP:07:14-16
03/09/92
0.181
0.292
32.0
5 4
ENP:07:16-18
03/09/92
0.193
0.304
31 8
4.2
ENP:07:18-20
03/09/92
0.149
0.252
34.5
ENP:07:20-22
03/09/92
0217
0.328
ENP:07:22-24
03/09/92
0.247
0367
ENP:07:24-26
03/09/92
0.269
0.394
19.7
ENP:07:26-28
03/09/92
0.390
0,566
ENP:07:28-30
03/09/92
0.362
0.529
ENP:07:30-32
03/09/92
0.383
0576
159
10.4
ENP:07:32-34
03/09/92
0.367
0.530
ENP:07:34-36
03/09/92
0.445
0.668
ENP:07:36-38
03/09/92
0.405
0593
164
ENP:07:3840
03/09/92
0.386
0582
ENP:07:40-42
03/09/92
0.377
0533
ENP:07:4244
03/09/92
0 423
0.621
16 1
Organic deposition Total
Carbon Pb-210 Cs-137 period rate Mercury
% pCi/g pCi/g year g/cmA2-yr ng/g
14
BDL
10
BDL
BDL
BDL
BDL
BDL
BDL
BDL
BDL
10
18.2
5.19
0.72
1992
0.079
100
5.19
0.72
1988
0.070
48
10.9
3.69
074
1979
0.073
35
14.4
4.26
1.19
1967
0 045
43
17.8
2.11
1.53
1949
0.050
31
14.9
1.21
0.84
1927
0.044
35
22.5
0.48
1.14
1903
0.054
NA
26.6
0.20
0 49
1890
0.084
48
27.6
0.15
0.31
1882
0.088
29
0.48
0.29
1874
0.022
45
0.14
0.03
1850
0.020
30
0.15
19
0.09
32
007
14
18
5.5 12
10
10
16
22
31
12
O'
u>

Table A.4. (cont'd)
Solids
Bulk
Total
Inorganic
Sampling
dry wt
Density
Carbon
Carbon
Sample I D.
Date
g/g
g/cmA3
%
%
ENP:07:44-46
03/09/92
0.420
0.641
ENP:07:46-48
03/09/92
0.488
0.726
ENP:07:48-50
03/09/92
0.405
0.595
15.7
ENP:07:50-52
03/09/92
0.418
0.606
ENP:07:52-54
03/09/92
0.461
0.713
ENP:07:54-56
03/09/92
0.474
0.737
15.8
10.4
ENP:07:56-58
03/09/92
0.481
0.714
ENP:07:58-60
03/09/92
0.472
0.699
ENP:07:60-61
03/09/92
0.437
0.654
17.4
ENP:08A:00-02
03/09/92
0.155
0.148
ENP:08A:04-06
03/09/92
0.148
0 142
ENP:08A:06-08
03/09/92
0.164
0.159
ENP:08A:08-10
03/09/92
0.169
0.165
ENP:08A: 10-12
03/09/92
0.150
0.146
ENP:08A:12-14
03/09/92
0.168
0.164
ENP:08A:16-18
03/09/92
0.135
0,130
ENP:08A: 18-20
03/09/92
0.149
0.146
ENP:08A:20-22
03/09/92
0.152
0.143
ENP:08A:22-24
03/09/92
0.124
0.116
ENP:08A:24-26
03/09/92
0.130
0.124
ENP:08A:26-28
03/09/92
0.110
0.105
ENP:08A: 30-32
03/09/92
0.151
0.143
ENP:08A:32-34
03/09/92
0.174
0.165
ENP:08A:34-36
03/09/92
0.150
0.144
ENP:08A:36-39
03/09/92
0.165
0.160
ENP:08B:00-02
03/09/92
0.126
0.134
ENP:08B:04-06
03/09/92
0.095
0.098
ENP:08B:06-08
03/09/92
0.113
0.125
ENP:08B:08-10
03/09/92
0.144
0.141
ENP:08B:10-12
03/09/92
0.139
0.136
ENP:08B:12-14
03/09/92
0.132
0.126
ENP:08B:16-18
03/09/92
0 144
0.137
ENP:08B: 18-20
03/09/92
0.140
0.138
ENP:08B:20-22
03/09/92
0.164
0.160
Organic
Carbon
91
85
89
94
86
75
93
57
67
93
77
116
112
65
98
98
118
171
157
174
116
89
103
116
81
deposition Total
Pb-210 Cs-137 period rate Mercury
pCi/g pCi/g year g/cmA2-vr ng/g
23
BDL
11
13
10
10
12
10
20
ON
4^

Table A.4. (cont'd)
Solids
Bulk
Total
Inorganic
Organic
deposition
Total
Sampling
dry wt
Density
Carbon
Carbon
Carbon
Pb-210
Cs-137
period rate
Mercury
Sample I D.
ENP:08B:22-24
Date
03/09/92
gl&
0.145
g/cmA3
0.140
%
%
%
pCi/g
pC]/g
vear g/cmA2-vr
81
ENP:08B:24-26
03/09/92
0.118
0.116
125
ENP:08B:26-28
03/09/92
0.136
0.129
108
ENP:08B:30-32
03/09/92
0.136
0.125
120
ENP:08B:32-34
03/09/92
0.150
0.140
118
ENP:09:00-02
03/10/92
0.156
0.157
10.23
0.53
1992
0.062
97
ENP:09:02-04
03/10/92
0.165
0.161
40.5
10.79
066
1986
0.049
52
ENP:09:04-06
03/10/92
0.149
0.146
8.97
0 50
1979
0.047
115
ENP:09:06-08
03/10/92
0.146
0.146
9 49
0 44
1972
0.036
87
ENP:09:08-10
03/10/92
0.163
0.160
9.20
0.39
1963
0.028
85
ENP:09:10-12
03/10/92
0.149
0.144
7.82
0.78
1948
0.021
86
ENP:09:12-14
03/10/92
0,137
0 132
5 27
046
1930
0017
79
ENP: 09:14-16
03/10/92
0 173
0 162
1.62
0.15
1909
0.030
56
ENP:09:16-18
03/10/92
0.144
0 136
1.83
0 13
1896
0.017
BDL
ENP:09:18-20
03/10/92
0.167
0.158
40.6
1.22
0.13
1874
0.013
13
ENP:09:20-22
03/10/92
0.160
0 153
0.44
0.18
1831
0.010
BDL
ENP:09:22-24
03/10/92
0.176
0.165
BDL
ENP:09:24-26
03/10/92
0.137
0 137
BDL
ENP:09:26-28
03/10/92
0.136
0 130
BDL
ENP:09:28-30
03/10/92
0.121
0.116
BDL
ENP:09:30-32
03/10/92
0.130
0.117
BDL
ENP:09:32-34
03/10/92
0,143
0.149
BDL
ENP:09:34-36
03/10/92
0.128
0.131
BDL
ENP:09:36-38
03/10/92
0.138
0.146
BDL
ENP:09:3840
03/10/92
0.158
0.153
BDL
ENP:09:40^t2
03/10/92
0.153
0.141
BDL
ENP:09:42-44
03/10/92
0.114
0.116
BDL
ENP:09:44-46
03/10/92
0.131
0.124
38.9
BDL
ENP:09:46-t8
03/10/92
0.127
0.104
BDL
ENP:10:00-02
03/10/92
0.088
0.095
49
ENP: 10:02-04
03/10/92
0,103
0 109
42
ENP: 10:04-06
03/10/92
0.107
0.116
41
ENP: 10:06-08
03/10/92
0.115
0.122
28
ENP: 10:08-10
03/10/92
0.114
0.121
38

Table A.4. (cont'd)
Solids
Bulk
Total
Inorganic
Sampling
dry wt
Density
Carbon
Carbon
Sample I D
Date
&1&
g/cmA3
%
%
ENP:10:10-12
03/10/92
0.103
0.109
ENP:10:12-14
03/10/92
0.108
0.119
ENP:10:14-16
03/10/92
0.127
0.142
ENP:10:16-18
03/10/92
0.145
0.157
ENP:10:18-20
03/10/92
0.161
0.176
ENP:10:20-22
03/10/92
0.148
0.155
ENP: 10:22-24
03/10/92
0.128
0 124
ENP:10:24-26
03/10/92
0.214
0.254
ENP:10:26-28
03/10/92
0.262
0301
ENP: 10:28-30
03/10/92
0.188
0.208
ENP: 10:30-32
03/10/92
0.161
0.166
ENP: 10:32-34
03/10/92
0.164
0.169
ENP: 10:34-36
03/10/92
0.157
0.172
ENP: 10:36-38
03/10/92
0.182
0.187
ENP: 10:38-40
03/10/92
0.399
0.486
ENP: 11:00-02
03/10/92
0.347
0.214
ENP: 11:02-04
03/10/92
0.354
0218
43.2
ENP: 11:04-06
03/10/92
0.347
0.213
ENP: 11:06-08
03/10/92
0.299
0.176
ENP: 11:08-10
03/10/92
0.317
0.183
ENP:11:10-12
03/10/92
0.242
0.127
ENP:11:12-14
03/10/92
0.230
0 122
ENP:11:14-16
03/10/92
0.222
0 133
48.6
ENP:11:16-18
03/10/92
0.265
0.160
ENP:11:18-20
03/10/92
0.270
0.161
ENP: 11:20-22
03/10/92
0.265
0.154
ENP: 11:22-24
03/10/92
0.251
0.143
ENP: 11:24-26
03/10/92
0.278
0.159
ENP: 11:26-28
03/10/92
0.268
0.162
ENP: 11:28-30
03/10/92
0.298
0.198
31.3
ENP: 11:30-32
03/10/92
0 393
0.341
ENP: 12:00-02
03/10/92
0.024
0.023
ENP:12:02-04
03/10/92
0.125
0.123
41.9
ENP: 12:04-06
03/10/92
0.090
0086
Organic deposition Total
Carbon Pb-210 Cs-137 period rate Mercury
% pCi/g pCi/g year g/cmA2-vr ng/g
42
30
25
22
34
51
34
30
29
17
60
59
48
53
BDL
10.60
1.82
1992
0.060
38
11.80
2.67
1984
0.042
40
12.60
2.64
1971
0.027
31
7.51
3.70
1949
0.023
25
3.57
2.26
1928
0.025
30
2.74
2.59
1908
0017
40
1 42
1.28
1888
0.018
23
1.04
0.47
1871
0.014
29
0.60
0.51
1842
0.010
32
28
28
30
19
57
70
33
10 62
1.07
1992
0.048
170
12 34
1.65
1991
0.040
86
13.70
2.38
1984
0.029
95
O'
O'

Table A.4. (cont'd)
Solids
Bulk
Total
Inorganic
Organic
deposition
Total
Sampling
dry wt
Density
Carbon
Carbon
Carbon
Pb-210
Cs-137
period
rate
Mercury
Sample I D.
Date
sis,
e/cmA3
%
%
%
pCi/g
pCi/fi
year
e/cmA2-vr
nsls
ENP: 12:06-08
03/10/92
0.104
0.102
13.71
3.59
1978
0.023
94
ENP: 12:08-10
03/10/92
0.128
0.129
42.1
8.47
3 26
1967
0.028
68
ENP:12:10-12
03/10/92
0.126
0 126
11.88
5.45
1956
0.014
41
ENP:12:12-14
03/10/92
0.119
0.120
1.68
0.70
1930
0.044
35
ENP: 12:14-16
03/10/92
0.109
0.111
3.02
0.37
1924
0.020
18
ENP:12:16-18
03/10/92
0.109
0.111
2.49
060
1911
0.016
18
ENP:12:18-20
03/10/92
0.204
0.221
1.68
0.63
1893
0.014
15
ENP: 12:20-22
03/10/92
0.302
0.336
0.00
0.05
BDL
ENP: 12:22-24
03/10/92
0.241
0.262
BDL
ENP:12:24-26
03/10/92
0.319
0.359
BDL
ENP: 12:26-28
03/10/92
0.362
0.438
BDL
ENP: 12:28-30
03/10/92
0.482
0609
BDL
ENP: 12:30-32
03/10/92
0.431
0.559
10
ENP:12:32-34
03/10/92
0.528
0.733
12
ENP:TS1:00-02
01/07/92
0.220
0.118
23.9
6.2
17.7
10.84
0.85
1992
0.061
39
ENP:TS1:02-04
01/07/92
0.305
0.151
20.9
7.5
13 3
7.73
0.86
1988
0.075
62
ENP:TS1:04-06
01/07/92
0.272
0 177
21.1
7.7
13.4
9 38
0.70
1984
0.054
55
ENP:TSl:06-08
01/07/92
0.341
0371
19.2
5.02
0.74
1976
0.080
48
ENP:TS1:08-10
01/07/92
0.406
0471
17.4
3.16
0.25
1965
0.090
28
ENP:TS1:10-12
01/07/92
0.435
0.542
16.7
9.4
7.3
1.81
0.22
1953
0.107
26
ENP:TS1:12-14
01/07/92
0.402
0.491
17.0
1.49
0.19
1940
0.089
41
ENP:TS1:14-16
01/07/92
0.357
0.454
18.2
6.1
12.0
1.19
0.27
1927
0.073
53
ENP:TS1:16-18
01/07/92
0.379
0.443
16.7
0.85
0.08
1911
0.062
43
ENP:TS1:18-20
01/07/92
0.499
0.561
14.1
9.2
4.9
050
0 26
1892
0.059
22
ENP:TS1:20-22
01/07/92
0.552
0766
0 25
0.10
1863
0.048
29
ENP:TSl:22-24
01/07/92
0.553
0 764
000
HIM.
31
ENP:TSl:24-26
01/07/92
0.548
0.693
12.6
0.09
14
ENP:TSl:26-28
01/07/92
0.628
0.938
0.04
NA
ENP:TS1:28-30
01/07/92
0.602
0.833
001
16
ENP:TS 1:30-32
01/07/92
0.604
0 771
119
8.3
3.6
16
ENP:TS2:00-02
01/07/92
0241
0 122
22.0
7.5
14 5
8.38
0 56
1992
0.069
36
ENP:TS2:02-04
01/07/92
0.373
0.274
17.8
9.3
8.5
6.97
0.65
1988
0.074
34
ENP:TS2:04-06
01/07/92
0.388
0.336
16.2
10.0
6.2
4.40
0.44
1980
0.090
NA
ENP:TS2:06-08
01/07/92
0.429
0447
15.7
2.42
046
1971
0.125
34

Table A.4. (cont'd)
Solids
Bulk
Total
Inorganic
Sampling
dry wt
Density
Carbon
Carbon
Samóle I D.
Date
g/g
g/cmA3
%
%
ENP:TS2:08-10
01/07/92
0.425
0.410
16.1
ENP:TS2:10-12
01/07/92
0.433
0413
16.0
ENP:TS2:12-14
01/07/92
0.439
0424
15.8
ENP:TS2:14-16
01/07/92
0.439
0.442
15.8
9.5
ENP:TS2:16-18
01/07/92
0.435
0.454
16.0
ENP:TS2:18-20
01/07/92
0.409
0.423
16.7
8.8
ENP:TS2:20-22
01/07/92
0.447
0.503
ENP:TS2:22-24
01/07/92
0.405
0.472
ENP:TS2:24-26
01/07/92
0.432
0.503
13.3
13.2
ENP:TS2:26-28
01/07/92
0.470
0625
ENP:TS2:28-30
01/07/92
0.518
0.590
ENP:TS2:30-32
01/07/92
0.540
0.790
13.3
9.1
BDL, below detection limit
NA, analysis not available
Organic
deposition
Total
Carbon
Pb-210
Cs-137
period
rate
Mercury
%
pCi/g
PCi/g
year
e/cmA2-vr
ng/g
2.06
0.30
1963
0.114
34
0.96
0.00
1955
0.190
29
0.85
0.14
1950
0.186
29
63
1.29
BDL
1945
0.105
20
0.77
0.20
1936
0.130
29
7.2
0.90
BDL
1928
0.087
34
1.14
0.06
1916
0.048
28
041
BDL
1882
0.046
41
0.1
022
BDL
1850
0.031
29
000
BDL
31
0.12
24
42
0.07
23
ON
00

Table A.5. Sediment and Mercury Analyses - Stormwater Treatment Area
Solids
Bulk
Total
Inorg.
Organic
deposition
Total
Sampling
dry wt
Density
Carbon
Carbon
Carbon
Pb-210
Cs-137
period rate
Mercury
Sample I D.
Date
SJ&
a/cmA3
%
%
%
pCi/fi
pC.i/g
year g/cmA2-vr
STA:43:00-02
07/15/92
0.189
0.203
47
STA:43:02-04
07/15/92
0.228
0248
55
STA:43:04-06
07/15/92
0.212
0230
59
STA:43:06-08
07/15/92
0.251
0272
58
STA:44:GRAB
07/15/92
0.304
0.269
70
STA:45:00-02
07/15/92
0.149
0.161
47
STA:45:02-04
07/15/92
0 209
0224
51
STA:45:04-06
07/15/92
0.228
0.250
31
STA:45:06-08
07/15/92
0.210
0228
33
STA:45:08-10
07/15/92
0.195
0.204
31
STA:45:10-12
07/15/92
0.139
0.146
70
STA:45:12-14
07/15/92
0.156
0.165
75
STA:45:14-17
07/15/92
0.120
0 129
50
ST A:46:00-02
07/15/92
0.145
0.151
93
STA:46:02-04
07/15/92
0.171
0.178
74
STA:46:04-06
07/15/92
0.229
0.248
61
STA:46:06-08
07/15/92
0.244
0.275
44
STA:46:08-10
07/15/92
0.235
0.254
46
STA: 46:10-12
07/15/92
0.235
0254
54
STA: 46:12-14
07/15/92
0.237
0.263
65
STA:46:14-16
07/15/92
0.245
0263
43
STA:46:16-18
07/15/92
0.269
0.294
37
STA:46:18-20
07/15/92
0.245
0.272
40
STA:47:A
07/15/92
0.294
0.262
30
STA:47:B
07/15/92
0.294
0287
33
STA:47:C
07/15/92
0.268
0.264
26
STA:47:D
07/15/92
0.270
0.252
29
STA:47:E
07/15/92
0.201
0.222
30
BDL, below detection limit
NA, analysis not available

Table A.6. Sediment and Mercury Analyses - Savannas
Solids
Bulk
Total
Inorg.
Organic
deposition
Total
Sampling
dry wt
Density
Carbon
Carbon
Carbon
Pb-210
Cs-137
period
rate
Mercury
Sample I D.
Date
g/g
g/cmA3
%
%
%
pCi/g
pCi/g
year
g/cmA2-yr
ng/g
SAV:48:00-01
01/19/93
0.130
0.129
17.05
5.97
1992
0.029
118
SAV:48:01-02
01/19/93
0.179
0.194
13.47
2.56
1987
0.032
128
SAV:48:02-03
01/19/93
0.175
0.180
11.03
2.55
1980
0.031
87
SAV:48:03-04
01/19/93
0.182
0.181
10.14
2,21
1974
0.028
137
SAV:48:04-05
01/19/93
0.183
0.185
7.19
1.99
1967
0032
109
SAV:48:05-06
01/19/93
0224
0.233
3.26
1.10
1960
0.057
88
SAV:48:06-08
01/19/93
0.242
0.271
363
0 81
1956
0.045
70
SAV:48:08-10
01/19/93
0.376
0.447
0.95
0.22
1941
0.106
28
SAV:48:10-12
01/19/93
0.296
0.324
1.37
0.26
1931
0.054
30
SAV:48:12-14
01/19/93
0.287
0 313
2.02
0.70
1916
0.023
42
SAV:48:14-16
01/19/93
0.296
0.329
0.37
0.68
1858
0.020
45
SAV:48:16-18
01/19/93
0.489
0.642
16
SAV:48:18-20
01/19/93
0.621
0.920
15
SAV:48:20-22
01/19/93
0.697
1.082
9
SAV:49:00-01
01/19/93
0.132
0.138
19.14
6.52
1992
0.022
115
SAV:49:01-02
01/19/93
0.204
0.215
12.75
4.97
1985
0.027
139
SAV:49:02-03
01/19/93
0.223
0.234
9.85
2,78
1976
0.026
154
SAV:49:03-04
01/19/93
0.282
0321
5.36
1.07
1965
0.034
85
SAV:49:04-05
01/19/93
0.212
0.223
2.73
1.64
1954
0.047
113
SAV:49:05-06
01/19/93
0.162
0.164
2.61
1.04
1948
0.042
113
SAV:49:06-07
01/19/93
0.166
0.168
3.20
1.63
1944
0030
120
SAV:49:07-08
01/19/93
0.164
0.163
1.80
1.25
1938
0 044
85
SAV:49:08-09
01/19/93
0.152
0.151
2 80
1.53
1934
0.025
110
SAV:49:09-11
01/19/93
0 132
0.130
2.19
1.18
1927
0.026
83
SAV:49:11-12
01/19/93
0.156
0.158
2.09
0.96
1915
0.018
52
SAV:49:12-14
01/19/93
0.144
0.141
1.81
0.75
1905
0.015
66
SAV:49:14-16
01/19/93
0.125
0.128
1.52
0.19
1878
0.008
65
SAV:49:16-18
01/19/93
0.170
0,168
73
SAV:49:18-20
01/19/93
0.350
0380
43
SAV:49:20-22
01/19/93
0.373
0.435
25
SAV:49:22-24
01/19/93
0.333
0,375
51
SAV:49:24-25
01/19/93
0.382
0.413
52
SAV:50:GRAB
01/27/93
0.240
0.094
52
SAV:53:00-02
01/27/93
0.133
0.137
68

Table A.6. (cont'd)
Solids
Bulk
Total
Inorg
Sampling
dry wt
Density
Carbon
Carbon
Sample I D
Date
g/g
g/cmA3
%
%
SAV:53:02-04
01/27/93
0.326
0.373
SAV:53:04-06
01/27/93
0.360
0417
SAV:53:06-08
01/27/93
0.356
0.405
SAV:53:08-10
01/27/93
0.391
0.469
SAV:53:10-12
01/27/93
0.512
0.685
SAV:53:12-14
01/27/93
0.587
0.863
SAV:53:14-16
01/27/93
0,630
0934
SAV:53:16-18
01/27/93
0.670
1.046
SAV: 53:18-20
01/27/93
0.723
1.230
SAV:53:20-22
01/27/93
0.745
1.297
SAV:53:22-24
01/27/93
0.750
1.313
SAV:53:24-25
01/27/93
0.749
1.317
SAV:54:00-04
01/27/93
0.809
0.638
SAV:54:04-08
01/27/93
0.787
0.841
SAV:54:08-12
01/27/93
0.824
0.748
BDL, below detection limit
NA, analysis not available
Organic
Carbon
%
50
53
55
30
27
25
18
14
11
13
15
deposition Total
Pb-210 Cs-137 period rate Mercury
pCi/g pCi/g year g/cmA2-yr ng/g
5
3
2

Table A. 7.
Sediment and Mercury Analyses - Okefenokee Swamp
Solids
Bulk
Total Inorg Organic
deposition
Total
Sampling
dry wt
Density
Carbon Carbon Carbon Pb-210
Cs-137
period rate
Mercury
Sample I D.
Date
g/g
e/cmA3
% % % pCi/g
pCi/fi
year g/cmA2-vr
OKE:56:00-02
02/15/93
0.050
0.049
154
OKE:56:02-04
02/15/93
0.060
0.061
124
OKE:56:04-06
02/15/93
0.069
0070
93
OKE:56:06-08
02/15/93
0.068
0068
108
OKE:56:08-10
02/15/93
0.061
0.062
82
OKE:56:10-12
02/15/93
0.071
0.070
68
OKE:56:12-14
02/15/93
0.067
0.066
57
OKE:56:14-16
02/15/93
0.065
0.062
74
OKE:56:16-18
02/15/93
0.060
0.058
85
OKE:56:18-20
02/15/93
0.091
0.091
42
OKE:56:20-22
02/15/93
0.083
0.084
30
OKE:57:00-02
02/15/93
0.091
0.090
80
OKE:57:02-04
02/15/93
0.088
0.089
112
OKE:57:04-06
02/15/93
0.076
0.075
94
OKE:57:06-08
02/15/93
0.077
0.077
96
OKE:57:08-10
02/15/93
0.083
0084
77
OKE:57:10-12
02/15/93
0.086
0.086
74
OKE:57:12-14
02/15/93
0.087
0.086
74
OKE:57:14-16
02/15/93
0.090
0.087
82
OKE:57:16-18
02/15/93
0.088
0,085
80
OKE:57:18-20
02/15/93
0.092
0.090
70
OKE:57:20-22
02/15/93
0.098
0.096
64
OKE:57:22-24
02/15/93
0.099
0.097
62
OKE:58:00-02
02/15/93
0.049
0.047
103
OKE:58:02-04
02/15/93
0.064
0.068
82
OKE:58:04-06
02/15/93
0 090
0.099
126
OKE:58:06-08
02/15/93
0.107
0.107
82
OKE:58:08-10
02/15/93
0.094
0.098
78
OKE:58:IO-12
02/15/93
0.108
0.111
59
OKE:58:12-14
02/15/93
0.095
0.095
53
OKE:58:14-16
02/15/93
0.090
0.091
57
OKE:58:16-18
02/15/93
0.082
0.083
77
OKE:58:18-20
02/15/93
0.102
0.106
61
BDL, below detection limit
NA, analysis not available

173
Table A.8. Metal Analyses - Water Conservation Area 1
Sampling
Cd
Cr
Cu
Fe
Ni
Pb
Zn
Sample I D.
Date
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
WCA1:01:00-02
06/18/92
3
6
23
2400
BDL
30
43
WCAl:01:02-04
06/18/92
3
BDL
16
1550
7
44
35
WCAl:01:04-06
06/18/92
3
BDL
23
1550
BDL
34
24
WCAl:01:06-08
06/18/92
3
BDL
21
1550
BDL
34
16
WCA1:01:08-10
06/18/92
3
BDL
18
2010
BDL
44
17
WCA1:01:10-12
06/18/92
BDL
BDL
21
2020
BDL
44
20
WCA1:01:12-14
06/18/92
BDL
BDL
18
2470
BDL
44
19
WCA1:01:14-16
06/18/92
BDL
BDL
20
2240
BDL
39
19
WCA1:01:16-18
06/18/92
BDL
BDL
26
3400
BDL
44
21
WCAl:01:18-20
06/18/92
BDL
BDL
16
3400
BDL
35
12
WCAl:01:20-22
06/18/92
WCAl:01:22-24
06/18/92
WCAl:01:24-26
06/18/92
BDL
BDL
7
2010
BDL
35
6
WCA1:01:26-28
06/18/92
WCAl:01:28-30
06/18/92
WCAl:01:30-32
06/18/92
BDL
BDL
5
2010
BDL
25
2
WCAl:01:32-34
06/18/92
WCAl:01:34-36
06/18/92
WCAl:01:36-38
06/18/92
BDL
BDL
5
500
BDL
16
5
WCAl:01:38-40
06/18/92
WCAl:01:40-42
06/18/92
WCAl:01:42-44
06/18/92
BDL
BDL
8
437
BDL
16
5
WCAl:01:44-46
06/18/92
WCA1:01:46-48
06/18/92
WCAl:01:48-50
06/18/92
BDL
BDL
BDL
487
BDL
BDL
BDL
WCAL0L5O-52
06/18/92
WCAl:01:52-54
06/18/92
WCA1:01:54-56
06/18/92
BDL
BDL
13
542
BDL
BDL
5
WCAl:01:56-58
06/18/92
WCA1:01:58-60
06/18/92
WCA1:01:60-62
06/18/92
BDL
BDL
BDL
778
BDL
BDL
3
WCAl:01:62-64
06/18/92
WCA 1:01:64-66
06/18/92
WCAl:01:66-68
06/18/92
BDL
BDL
BDL
836
BDL
BDL
BDL
WCAl:01:68-70
06/18/92
WCAl:01:70-72
06/18/92
WCAl:01:72-74
06/18/92
3
BDL
BDL
933
BDL
25
6
WCAl:01:74-76
06/18/92
WCAl:01:76-78
06/18/92
WCAl:01:78-80
06/18/92
WCAl:01:80-82
06/18/92
WCA1:01:82-84
06/18/92
WCAl:01:84-86
06/18/92
WCAl:01:86-88
06/18/92
WCAl:01:88-90
06/18/92
WCAl:01:90-92
06/18/92
BDL
BDL
BDL
679
BDL
16
BDL
WCAl:01:92-94
06/18/92
WCAl:01:94-96
06/18/92
WCAl:01:96-98
06/18/92
BDL
BDL
BDL
850
BDL
BDL
8
WCA1:01:98-100
06/18/92
WCA1:01:100-102
06/18/92
WCA1:01:102-103
06/18/92
BDL
BDL
BDL
807
BDL
16
15

Table A.8. (cont'd)
Sampling
Cd
Cr
Cu
Fe
Ni
Pb
Zn
Sample I D.
Date
mg/kg
mg/ke
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
WCAl:35:00-02
07/15/92
BDL
BDL
26
2910
BDL
34
35
WCA1-.35:02-04
07/15/92
2
BDL
25
2840
BDL
42
21
WCAl:35:04-06
07/15/92
BDL
BDL
16
2890
BDL
34
17
WCAl:35:06-08
07/15/92
BDL
BDL
11
2400
BDL
33
19
WCA1:35:08-10
07/15/92
BDL
BDL
10
2000
BDL
25
12
WCA1:35:10-12
07/15/92
BDL
BDL
12
2000
BDL
16
11
WCA1:35:12-14
07/15/92
BDL
BDL
10
1430
BDL
16
7
WCA1:35:14-16
07/15/92
BDL
BDL
10
1230
BDL
BDL
5
WCA1:35:16-18
07/15/92
BDL
BDL
5
1110
BDL
BDL
5
WCA1:35:18-20
07/15/92
BDL
BDL
7
1290
BDL
BDL
4
WCAl:35:20-22
07/15/92
WCA1:35:22-24
07/15/92
WCA1:35:24-26
07/15/92
BDL
BDL
7
1300
BDL
BDL
2
WCAl:35:26-28
07/15/92
WCAl:35:28-30
07/15/92
WCAl:35:30-32
07/15/92
BDL
BDL
BDL
1390
BDL
BDL
1
WCAl:35:32-34
07/15/92
WCAl:35:34-36
07/15/92
WCA1:36:00-02
07/15/92
BDL
7
15
2470
BDL
89
43
WCAl:36:02-04
07/15/92
BDL
BDL
16
1370
BDL
80
40
WCAl:36:04-06
07/15/92
BDL
BDL
10
1100
BDL
62
29
WCA1:36:06-08
07/15/92
BDL
BDL
15
1180
BDL
53
31
WCA1.36:08-10
07/15/92
BDL
BDL
BDL
1060
BDL
25
13
WCA1:36:10-12
07/15/92
BDL
BDL
BDL
1060
BDL
16
7
WCA1:36:12-14
07/15/92
BDL
BDL
BDL
1070
BDL
BDL
7
WCA1:36:14-16
07/15/92
BDL
BDL
12
1040
BDL
BDL
5
WCA1:36:16-18
07/15/92
BDL
BDL
BDL
1060
BDL
BDL
2
WCA1:36:18-20
07/15/92
BDL
BDL
BDL
1000
BDL
BDL
BDL
WCAl:36:20-22
07/15/92
WCAl:36:22-24
07/15/92
WCAl:36:24-26
07/15/92
WCAl:36:26-28
07/15/92
BDL
BDL
BDL
785
BDL
BDL
BDL
WCAl:36:28-30
07/15/92
WCAl:36:30-32
07/15/92
WCAl:36:32-34
07/15/92
BDL
BDL
BDL
704
BDL
BDL
BDL
WCAl:36:34-36
07/15/92
WCA1:36:36-38
07/15/92
WCAl:36:38-40
07/15/92
BDL
BDL
BDL
982
BDL
BDL
2
WCAl:36:40-42
07/15/92
BDL
BDL
BDL
1030
BDL
BDL
6
WCA1:36:42^14
07/15/92
BDL
BDL
5
1180
BDL
BDL
2
WCA1:36:44-46
07/15/92
WCAl:36:46-48
07/15/92
WCAl:36:48-50
07/15/92
WCAl:36:50-52
07/15/92
BDL
BDL
BDL
1400
BDL
BDL
6
WCAl:37:00-02
07/15/92
WCA1:37:02-04
07/15/92
WCA1:37:04-06
07/15/92
BDL
BDL
8
2890
BDL
34
30
WCAl:37:06-08
07/15/92
BDL
BDL
13
1430
BDL
61
36
WCA1:37:08-10
07/15/92
BDL
BDL
13
1270
BDL
52
32
WCA1:37:10-12
07/15/92
BDL
BDL
11
1200
BDL
77
25
WCA1:37:12-14
07/15/92
BDL
BDL
7
975
BDL
20
10
WCA1:37:14-16
07/15/92
BDL
BDL
5
1080
BDL
25
5

175
Table A.8. (cont'd)
Sampling
Cd
Cr
Sample I D.
Date
ingdcg
mg/kg
WCA1:37:16-18
07/15/92
BDL
BDL
WCAl:37:18-20
07/15/92
BDL
BDL
WCA1:37:20-22
07/15/92
WCA1:37:22-24
07/15/92
WCA1:37:24-26
07/15/92
BDL
BDL
WCA1:37:26-28
07/15/92
WCAl:37:28-30
07/15/92
WCAl:37:30-32
07/15/92
BDL
BDL
WCAl:37:32-34
07/15/92
WCA1:37:34-36
07/15/92
WCA1:37:36-38
07/15/92
BDL
BDL
WCA1:37:38^10
07/15/92
WCA1:37:40-42
07/15/92
WCAl:37:42-t4
07/15/92
BDL
BDL
WCA1:37:44-46
07/15/92
WCA1:37:46-48
07/15/92
WCA1:37:48-50
07/15/92
BDL
BDL
WCAl:37:50-52
07/15/92
WCAl:37:52-54
07/15/92
WCA 1:37:54-56
07/15/92
BDL
BDL
WCAl:37:56-57
07/15/92
BDL, below detection limit
NA, analysis not available
Cu
mg/kg
BDL
5
Fe
1350
2410
Ni
mg/kg
BDL
BDL
Pb
mg/kg
BDL
BDL
Zn
mg/kg
6
4
5
1310
BDL
16
5
BDL
1930
BDL
BDL
3
BDL
2430
BDL
16
3
BDL
1970
BDL
BDL
2
5
1440
BDL
BDL
3
BDL
2920
BDL
BDL
BDL

176
Table A.9. Metal Analyses - Water Conservation Area 2
Sampling
Cd
Cr
Cu
Fe
Ni
Pb
Zn
Samóle I D.
Date
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
WCA2:26A:00-02
03/11/92
3
BDL
10
2920
BDL
62
17
WCA2:26A:02-04
03/11/92
2
7
li
3140
BDL
62
19
WCA2:26A: 04-06
03/11/92
BDL
BDL
10
3780
BDL
34
12
WCA2:26A: 06-08
03/11/92
BDL
BDL
BDL
3390
BDL
16
12
WCA2:26A: 08-10
03/11/92
BDL
BDL
BDL
3840
BDL
BDL
2
WCA2:26A: 10-12
03/11/92
BDL
BDL
BDL
3840
BDL
16
2
WCA2:26A. 12-14
03/11/92
BDL
BDL
5
2920
BDL
16
1
WCA2:26A:14-16
03/11/92
BDL
BDL
BDL
2910
BDL
BDL
2
WCA2:26A:16-18
03/11/92
BDL
BDL
BDL
2900
BDL
BDL
BDL
WCA2:26A: 18-20
03/11/92
BDL
BDL
BDL
2900
BDL
16
2
WCA2:26A:20-22
03/11/92
WCA2:26A:22-23
03/11/92
BDL
BDL
BDL
3160
BDL
BDL
8
WCA2:29:00-02
03/12/92
BDL
BDL
BDL
2370
BDL
33
33
WCA2:29:02-04
03/12/92
BDL
BDL
24
2390
BDL
25
24
WCA2:29:04-06
03/12/92
BDL
BDL
23
1910
BDL
33
26
WCA2:29:06-08
03/12/92
BDL
BDL
27
2380
BDL
24
26
WCA2:29:08-10
03/12/92
BDL
BDL
28
1060
BDL
25
26
WCA2:29:10-12
03/12/92
BDL
BDL
23
2810
BDL
33
29
WCA2:29:12-14
03/12/92
BDL
BDL
28
3300
BDL
43
26
WCA2:29:14-16
03/12/92
BDL
BDL
20
2370
BDL
33
21
WCA2:29:16-18
03/12/92
BDL
BDL
34
3340
BDL
43
33
WCA2:29:18-20
03/12/92
BDL
BDL
36
3700
BDL
33
22
WCA2:29:20-22
03/12/92
WCA2:29:22-24
03/12/92
WCA2:29:24-26
03/12/92
BDL
BDL
8
3260
BDL
BDL
9
WCA2:29:26-28
03/12/92
WCA2:29:28-30
03/12/92
WCA2:29:30-32
03/12/92
BDL
BDL
BDL
4580
BDL
BDL
16
WCA2:29:32-34
03/12/92
WCA2:29:34-36
03/12/92
WCA2:29:36-38
03/12/92
BDL
BDL
BDL
4560
BDL
BDL
2
WCA2:29:38-tO
03/12/92
WCA2:29:40-41
03/12/92
BDL
BDL
5
4220
BDL
BDL
6
WCA2:30:00-02
03/12/92
BDL
BDL
13
2020
BDL
35
22
WCA2:30:02-04
03/12/92
BDL
BDL
14
2760
BDL
35
21
WCA2:30:04-06
03/12/92
BDL
BDL
12
2110
BDL
27
18
WCA2:30:06-08
03/12/92
BDL
BDL
12
1780
BDL
23
18
WCA2:30:08-10
03/12/92
BDL
BDL
10
2460
BDL
25
20
WCA2:30:10-12
03/12/92
BDL
BDL
10
2450
BDL
25
16
WCA2:30:12-14
03/12/92
BDL
BDL
8
2440
BDL
25
13
WCA2:30:14-16
03/12/92
BDL
BDL
5
2450
BDL
16
6
WCA2:30:16-18
03/12/92
BDL
BDL
5
2020
BDL
16
7
WCA2:30:18-20
03/12/92
BDL
BDL
5
2010
BDL
16
5
WCA2:30:20-22
03/12/92
BDL
BDL
5
2000
BDL
BDL
5
WCA2:30:22-24
03/12/92
WCA2:30:24-26
03/12/92
WCA2:30:26-28
03/12/92
WCA2:30:28-30
03/12/92
BDL
BDL
BDL
2460
BDL
BDL
5
WCA2:30:30-32
03/12/92
WCA2:30:32-34
03/12/92
WCA2:30:34-36
03/12/92
BDL
BDL
BDL
2950
BDL
BDL
4
BDL, below detection limit
NA, analysis not available

177
Table A. 10. Metal Analyses - Water Conservation Area 3
Sampling
Cd
Cr
Cu
Fe
Ni
Pb
Zn
SamDle I D.
Date
mft/kg
mg/kg
mg/kg
mg/kg
mgZkg
mg/kg
mg/kg
WCA3:14:00-02
03/10/92
BDL
BDL
12
7910
5
45
69
WCA3:14:02-04
03/10/92
BDL
5
14
6420
BDL
55
45
WCA3:14:04-06
03/10/92
BDL
BDL
5
6600
BDL
62
18
WCA3:14:06-08
03/10/92
BDL
5
BDL
6680
BDL
26
30
WCA3:14:08-10
03/10/92
BDL
BDL
18
6560
BDL
25
17
WCA3:14:10-12
03/10/92
BDL
BDL
8
7130
BDL
16
6
WCA3:14:12-14
03/10/92
BDL
BDL
BDL
6120
BDL
43
9
WCA3:14:14-16
03/10/92
WCA3:14:16-18
03/10/92
BDL
BDL
BDL
10100
BDL
25
4
WCA3:14:18-20
03/10/92
BDL
BDL
10
11700
BDL
25
4
WCA3:14:20-22
03/10/92
WCA3:14:22-24
03/10/92
WCA3:14:24-26
03/10/92
4
7
7
4290
10
43
11
WCA3:14:26-28
03/10/92
WCA3:14:28-30
03/10/92
5
9
7
3145
19
52
4
WCA3:14:30-32
03/10/92
WCA3:14:32-34
03/10/92
5
7
8
2950
17
62
5
WCA3:17:00-02
03/10/92
BDL
BDL
8
10900
BDL
60
31
WCA3:17:02-04
03/10/92
BDL
BDL
8
8310
BDL
70
26
WCA3:17:04-06
03/10/92
BDL
BDL
10
8410
BDL
71
26
WCA3:17:06-08
03/10/92
BDL
BDL
7
BDL
BDL
42
8
WCA3:17:08-10
03/10/92
WCA3:17:10-12
03/10/92
BDL
BDL
7
7070
BDL
16
8
WCA3:17:12-14
03/10/92
BDL
BDL
5
6810
BDL
16
6
WCA3:17:14-16
03/10/92
BDL
BDL
5
8000
BDL
16
6
WCA3:17:16-18
03/10/92
BDL
BDL
7
8350
BDL
16
7
WCA3:18:00-02
03/11/92
WCA3:18:02-04
03/11/92
WCA3:18:04-06
03/11/92
BDL
BDL
8
5740
BDL
72
34
WCA3:18:06-08
03/11/92
BDL
BDL
8
6580
BDL
71
33
WCA3:18:08-10
03/11/92
BDL
BDL
7
6950
BDL
34
16
WCA3:18:10-12
03/11/92
BDL
BDL
5
6520
BDL
16
10
WCA3:18:12-14
03/11/92
BDL
BDL
8
4700
BDL
BDL
13
WCA3:18:14-16
03/11/92
BDL
BDL
7
6180
BDL
44
15
WCA3:18:16-18
03/11/92
BDL
BDL
7
5690
BDL
16
11
WCA3:18:18-20
03/11/92
BDL
BDL
5
5150
BDL
BDL
7
WCA3:18:20-22
03/11/92
WCA3:18:22-24
03/11/92
WCA3:18:24-26
03/11/92
BDL
BDL
5
6190
BDL
BDL
6
WCA3:18:26-28
03/11/92
WCA3:18:28-30
03/11/92
WCA3:18:30-32
03/11/92
BDL
BDL
7
6540
BDL
16
8
WCA3:18:32-34
03/11/92
WCA3:18:34-36
03/11/92
WCA3:18:36-38
03/11/92
WCA3:18:38-40
03/11/92
BDL
BDL
5
8010
BDL
BDL
4
WCA3:18:40-42
03/11/92
WCA3:18:42-44
03/11/92
BDL
7
5
8520
BDL
16
3
WCA3:18:44-46
03/11/92
WCA3:18:46-48
03/11/92
WCA3:18:48-50
03/11/92
WCA3:18:50-52
03/11/92
WCA3:18:52-53
03/11/92
BDL, below detection limit

178
Table A. 11. Metal Analyses - Everglades National Park
Sampling
Cd
Cr
Cu
Fe
Ni
Pb
Zn
Samóle I D.
Date
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
ENP:07:00-02
03/09/92
4
10
10
21800
10
53
24
ENP:07:02-04
03/09/92
ENP:07:04-06
03/09/92
4
10
8
4810
ii
62
4
ENP:07:06-08
03/09/92
4
10
8
6100
10
61
4
ENP:07:08-10
03/09/92
4
10
8
6990
10
52
4
ENP:07:10-12
03/09/92
4
12
8
6030
13
52
1
ENP:07:12-14
03/09/92
4
12
8
6530
10
52
2
ENP:07.14-16
03/09/92
3
10
7
7500
10
43
BDL
ENP:07:16-18
03/09/92
4
10
8
17200
7
44
23
ENP:07:18-20
03/09/92
4
10
7
10200
7
52
6
ENP:07:20-22
03/09/92
ENP:07:22-24
03/09/92
ENP:07:24-26
03/09/92
5
12
10
7590
11
62
8
ENP:07:26-28
03/09/92
ENP:07:28-30
03/09/92
ENP:07:30-32
03/09/92
4
12
10
2940
10
62
13
ENP:07:32-34
03/09/92
ENP:07:34-36
03/09/92
ENP:07:36-38
03/09/92
4
12
8
9710
10
61
BDL
ENP:07:38-40
03/09/92
ENP:07:40-42
03/09/92
ENP:07:4244
03/09/92
5
12
13
4730
14
70
6
ENP:07:44-46
03/09/92
ENP:07:46-48
03/09/92
ENP:07:48-50
03/09/92
ENP:07:50-52
03/09/92
ENP:07:52-54
03/09/92
ENP:07:54-56
03/09/92
5
12
10
3350
10
70
3
ENP:07:56-58
03/09/92
ENP:07:58-60
03/09/92
ENP:07:60-61
03/09/92
5
12
10
3390
14
53
6
ENP:TSl:00-02
01/07/92
4
15
15
7010
14
79
23
ENP:TS1:02-04
01/07/92
4
21
15
6510
16
91
23
ENP:TS 1:04-06
01/07/92
4
17
14
5950
16
95
21
ENP:TSl:06-08
01/07/92
4
9
12
7100
16
84
13
ENP:TS1:08-10
01/07/92
4
21
12
5830
16
76
8
ENP:TS1:10-12
01/07/92
4
19
13
5660
15
63
6
ENP:TS1:12-14
01/07/92
4
27
13
10500
17
61
8
ENP:TS1:14-16
01/07/92
4
52
12
19800
20
52
11
ENP:TS1:16-18
01/07/92
4
71
12
28400
20
52
9
ENP:TS1:18-20
01/07/92
4
32
10
11100
23
52
4
ENP:TSl:20-22
01/07/92
ENP:TSl:22-24
01/07/92
ENP:TS1:24-26
01/07/92
4
49
10
19300
20
61
6
EKP:TSl:26-28
01/07/92
ENP:TSl:28-30
01/07/92
ENP:TSl:30-32
01/07/92
4
44
8
19900
24
62
2
ENP:TS2:00-02
01/07/92
4
12
13
6140
10
71
23
ENP:TS2:02-04
01/07/92
4
15
15
3840
14
89
18
ENP:TS2:04-06
01/07/92
5
15
12
2910
17
80
8
ENP:TS2:06-08
01/07/92
5
17
13
4260
14
79
13
ENP:TS2:08-10
01/07/92
5
15
10
1090
17
71
4

Table A. 11. (cont'd)
Sampling
Cd
Cr
Sample I D.
Date
me/kg.
mgdcg
ENP:TS2:10-12
01/07/92
4
17
ENP:TS2:12-14
01/07/92
4
15
ENP:TS2:14-16
01/07/92
4
17
ENP:TS2:16-18
01/07/92
4
20
ENP:TS2:18-20
01/07/92
4
25
ENP:TS2:20-22
01/07/92
ENP:TS2:22-24
01/07/92
ENP:TS2:24-26
01/07/92
4
39
ENP:TS2:26-28
01/07/92
ENP:TS2:28-30
01/07/92
ENP:TS2:30-32
01/07/92
4
44
BDL, below detection limit
NA, analysis not available
Cu
Fe
Ni
Pb
Zn
mg/kg
mg/'kg
mft/kg
m.g/kg
mg/kg
ll
5600
16
61
4
26
5190
li
61
6
10
5220
5
62
1
12
6610
8
62
2
13
8880
8
62
3
13
13000
11
53
3
12
16200
BDL
61
13

BIOGRAPHICAL SKETCH
Brian Eugene Rood was bora on November 13, 1963, in Manchester, Connecticut,
to F. Eugene and Roberta Rood. He attended the University of Connecticut from 1981
through 1986 where he received Bachelor of Science degrees m chemistry and biology
He was recognized as an honors scholar by the University of Connecticut after completion
of his undergraduate research in chemical limnology.
Mr. Rood contmued his research in the field of limnology under the direction of
Dr. Edward S. Deevey, Jr., at the Florida Museum of Natural History, during which time
he received a Master of Science degree m zoology at the University of Florida. His
master's research studied the kinetics of immediate chemical oxygen demand m lakes.
After a period of employment at the Environmental Research Institute m Storrs,
Connecticut, he returned to the Department of Environmental Engineering Sciences at the
University of Florida where he has pursued his doctoral studies of the biogeochemistry
and paleolimnology of trace metals m soil/sediment of the Flonda Everglades and selected
wetlands.
180

I certify that I have read this study and that in my opinion it conforms to
acceptable standards of scholarly presentation and is fully adequate, in scone and quality,
as a dissertation for the degree of Doctor of Philosophy—^, _ / I
Jo'séph J Delfino, Chah
Professor of Environmental
Engineering Sciences
I certify that I have read this study and that m my opinion it conforms to
acceptable standards of scholarly presentation and is fully adequate, in scope and quality,
as a dissertation for the degree of Doctor of Philosophy.
George R. Best
Scientist of Environmental
Engineering Sciences
I certify that I have read this study and that in my opmion it conforms to
acceptable standards of scholarly presentation and is fully adequate, in scope and quality,
as a dissertation for the degree of Doctor of Philosophy.
William E. Bolch, Jr.
Professor of Environmental
Engineering Sciences
I certify that I have read this study and that m my opinion it conforms to
acceptable standards of scholarly presentation and is fully adequate, in scope and quality,
as a dissertation for the degree of Doctor of Philosophy.
/}.
Donald A. Graetz
Professor of Soil and Water Science
I certify that I have read this study and that m my opmion it conforms to
acceptable standards of scholarly presentation and is fully adequate, in scope and quality,
as a dissertation for the degree of Doctor of Philosophy
Frank G. Nordlie
Professor of Zoology

This dissertation was submitted to the Graduate Faculty of the College of
Engineering and to the Graduate School and was accepted as partial fulfillment of the
requirements for the degree of Doctor of Philosophy.
December, 1993
Dean, College of Engineering
Karen A. Holbrook
Dean, Graduate School

UNIVERSITY OF FLORIDA
I ill ill mu urn y ñ a c c
3 1262 08554 0465



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