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

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Spatial and temporal distribution of mercury and other metals in Florida Everglades and Savannas marsh soils
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Includes bibliographical references (leaves 124-142).
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by Brian Eugene Rood.
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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














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("4


t- On 0 ro ( oo 00
D 0\1 N Iv 0N O \t


U o

g o go P
5 4 V
(1) WMM


o-
.0 00
<'


c00 -N -C Cm


E -
E


0 -

C u





l~ o



00









.
- -







0 3


en Cu
CO
o m

0 an












* *








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