• TABLE OF CONTENTS
HIDE
 Copyright
 Title Page
 Acknowledgement
 Table of Contents
 List of Tables
 List of Figures
 Abstract
 Introduction
 Literature review
 Areas of investigation, sampling...
 Results and discussion
 Summary and conclusions
 Appendix
 Bibliography
 Biographical sketch














Title: Heavy metal interactions with natural organics in aquatic environments
CITATION THUMBNAILS PAGE IMAGE ZOOMABLE
Full Citation
STANDARD VIEW MARC VIEW
Permanent Link: http://ufdc.ufl.edu/UF00086027/00001
 Material Information
Title: Heavy metal interactions with natural organics in aquatic environments
Physical Description: xiii, 155 leaves : ill. ; 28 cm.
Language: English
Creator: Carriker, Neil Edward, 1948-
Publication Date: 1977
 Subjects
Subject: Heavy metals -- Environmental aspects   ( lcsh )
Sewage -- Analysis   ( lcsh )
Sewage -- Environmental aspects   ( lcsh )
Trace elements in water   ( lcsh )
Environmental Engineering Sciences thesis Ph. D
Dissertations, Academic -- Environmental Engineering Sciences -- UF
Genre: bibliography   ( marcgt )
non-fiction   ( marcgt )
 Notes
Thesis: Thesis--University of Florida.
Bibliography: Bibliography: leaves 148-154.
Statement of Responsibility: by Neil Edward Carriker.
General Note: Typescript.
General Note: Vita.
 Record Information
Bibliographic ID: UF00086027
Volume ID: VID00001
Source Institution: University of Florida
Rights Management: All rights reserved by the source institution and holding location.
Resource Identifier: aleph - 000209996
oclc - 04164650
notis - AAX6815

Table of Contents
    Copyright
        Copyright
    Title Page
        Page i
    Acknowledgement
        Page ii
        Page iii
    Table of Contents
        Page iv
        Page v
    List of Tables
        Page vi
        Page vii
    List of Figures
        Page viii
        Page ix
        Page x
        Page xi
    Abstract
        Page xii
        Page xiii
    Introduction
        Page 1
        Page 2
        Page 3
    Literature review
        Page 4
        Page 5
        Page 6
        Page 7
        Page 8
        Page 9
        Page 10
        Page 11
        Page 12
        Page 13
        Page 14
        Page 15
        Page 16
        Page 17
        Page 18
        Page 19
        Page 20
        Page 21
        Page 22
        Page 23
        Page 24
        Page 25
    Areas of investigation, sampling procedures, and analytical methods
        Page 26
        Page 27
        Page 28
        Page 29
        Page 30
        Page 31
        Page 32
        Page 33
        Page 34
        Page 35
        Page 36
        Page 37
        Page 38
        Page 39
        Page 40
        Page 41
        Page 42
        Page 43
        Page 44
        Page 45
        Page 46
        Page 47
        Page 48
    Results and discussion
        Page 49
        Page 50
        Page 51
        Page 52
        Page 53
        Page 54
        Page 55
        Page 56
        Page 57
        Page 58
        Page 59
        Page 60
        Page 61
        Page 62
        Page 63
        Page 64
        Page 65
        Page 66
        Page 67
        Page 68
        Page 69
        Page 70
        Page 71
        Page 72
        Page 73
        Page 74
        Page 75
        Page 76
        Page 77
        Page 78
        Page 79
        Page 80
        Page 81
        Page 82
        Page 83
        Page 84
        Page 85
        Page 86
        Page 87
        Page 88
        Page 89
        Page 90
        Page 91
        Page 92
        Page 93
        94
        Page 95
        Page 96
        Page 97
        Page 98
        Page 99
        Page 100
        Page 101
        Page 102
        Page 103
        Page 104
        Page 105
        Page 106
        Page 107
        Page 108
        Page 109
        Page 110
        Page 111
        Page 112
        Page 113
        Page 114
        Page 115
        Page 116
        Page 117
        Page 118
        Page 119
        Page 120
        Page 121
        Page 122
        Page 123
        Page 124
        Page 125
        Page 126
        Page 127
        Page 128
        Page 129
        Page 130
        Page 131
        Page 132
        Page 133
        Page 134
        Page 135
        Page 136
        Page 137
        Page 138
        Page 139
        Page 140
        Page 141
    Summary and conclusions
        Page 142
        Page 143
        Page 144
        Page 145
    Appendix
        Page 146
        Page 147
    Bibliography
        Page 148
        Page 149
        Page 150
        Page 151
        Page 152
        Page 153
        Page 154
    Biographical sketch
        Page 155
        Page 156
        Page 157
Full Text






U LJ UNIVERSITY of

UF FLORIDA
V;-' tr lld1tritn1 fi' I h' Calor Na tNi

Internet Distribution Consent Agreement

In reference to the following dissertation:

AUTHOR: Carriker, Neil
TITLE: Heavy metal interactions with natural organic in aquatic
environments / (record number. 209996)
PUBLICATION 1977
DATE:


I, ieA) FJ ,L A .1 as copyright holder for the
aforementioned dissertation, hereby grant specific and limited archive and distribution
rights to the Board of Trustees of the University of Florida and its agents. I authorize the
University of Florida to digitize and distribute the dissertation described above for
nonprofit, educational purposes via the Internet or successive technologies.
This is a non-exclusive grant of permissions for specific off-line and on-line uses for an
indefinite term. Off-line uses shall be limited to those specifically allowed by "Fair Use"
as prescribed by the terms of United States copyright legislation (cf, Title 17, U.S. Code)
as well as to the maintenance and preservation of a digital archive copy. Digitization
allows the University of Florida to generate image- and text-based versions as appropriate
and to provide and enhance access using search software.
This grant of permissions prohibits use of the digitized versions for commercial use or
profit.


Signature of Copyright Holder

Printed or Typed Name of Copyright Holder/Licensee


Personal information blurred



Date of Signature
















HEAVY METAL INTERACTIONS WITH NATURAL ORGANIC
IN AQUATIC ENVIRONMENTS










By

NEIL EDWARD CARRIKER


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



UNIVERSITY OF FLORIDA


1977














ACKNOWLEDGMENTS

I wish to express my sincere appreciation to the

chairman of my supervisory committee, Dr. P. L. Brezonik,

for his constructive suggestions during both the experi-

mental part of this work and the preparation of the manu-

script. I would also like to thank the other members of

my committee for their continued interest and suggestions

during these investigations.

This project was funded by a grant through the Center

for Wetlands jointly sponsored by the RANN Division of the

National Science Foundation and the Rockefeller Foundation.

I would like to thank Dr. H. T. Odum, who was primarily

responsible for securing these grants, and Dr. P. L.

Brezonik, who offered me the opportunity to participate in

this project.

Appreciation is also extended to Bob Klein, who pre-

ceded me in the heavy metals aspect of this project, to

Hugh Prentice, who assisted with the sediment coring and to

Woody Dierberg, who assisted with the routine sampling and

who provided much helpful assistance during the course of

these investigations. I also wish to thank Ms. Barbara

Smerage for a highly professional job of typing the manu-

script, with short notice and less than optimal conditions.








I especially wish to express my appreciation to my

wife for her continued patience and support during my

graduate studies, and to my mother, who has always en-

couraged me to seek excellence in all things.


iii
















TABLE OF CONTENTS

ACKNOWLEDGMENTS. . . . . .. ii

LIST OF TABLES . . . . .. vi

LIST OF FIGURES. .. . . . .viii

ABSTRACT . . . . . xii

CHAPTER

I. INTRODUCTION. . . . 1

II. LITERATURE REVIEW. . . . 4
A. Environmental Chemistry of Heavy Metals. 4
B. Analytical Methods Applicable to Heavy
Metal-Organic Interaction Studies . 9
1. Instrumental Methods for Measuring
Heavy Metal Concentrations . 9
2. Methods for Determining Molecular
Weight Distributions of Organics . 14
3. Methods for Observing Heavy Metal-
Organic Interactions. . .. 15
C. Investigation of Heavy Metals in Sewage. 19
D. Investigations of Heavy Metal-Natural
Organic Interactions . . ... .21

III. AREAS OF INVESTIGATION, SAMPLING PROCEDURES,
AND ANALYTICAL METHODS. . . .. 26
A. Field Studies. . . . .. .26
1. Site Description . . . 26
2. Routine Monitoring . . 28
3. Sediment Characterization and Metal
Profiles . . . 34
4. Plant Tissue Analysis. . . 37
B. Laboratory Studies. . . .. 38
1. Characterization of Natural and
Sewage Organics. ..... . 38
2. Heavy Metal Uptake by Duckweed . .42
3. Heavy Metal Release via Duckweed
Decomposition. . . ... 43
4. Effects of Model Organic Compounds on
DPASV Diagnostic Parameters and ISE
Response . . . . 44









5. Effects of Natural and Sewage Organics
on ASV Diagnostic Parameters and ISE
Response . . . ... .46
6. Determination of Complexation Capacity
of Model Organic Compounds and Natural
and Sewage Organics by Ion Exchange
Equilibrium. . . . ... 47

IV. RESULTS AND DISCUSSION . . ... .49
A. Field Studies. ............. .49
1. Routine Monitoring ......... 49
2. Sediment Characterization and Metal
Profiles . . . ... .57
3. Plant Tissue Analyses. . .. .78
B. Laboratory Studies . . ... .80
1. Characterization of Natural and
Sewage Organics. . . ... 80
2. Heavy Metal Uptake by Duckweed . 89
3. Heavy Metal Release via Duckweed
Decomposition. . . . .100
4. DPASV and ISE Investigations with


Model Compounds and Natural
Sewage Organics. . .
5. Ion Exchange Equilibrium
Investigations with Model
Compounds and Natural and
Sewage Organics. . .


V. SUMMARY AND CONCLUSIONS.

APPENDIX 1 . . .

BIBLIOGRAPHY . . .

BIOGRAPHICAL SKETCH. . .


and
. . .101



. . .128


. . . .142

. . . .146

. . . .148

. . . .155















LIST OF TABLES


Table Page

1 Typical concentrations of selected
heavy metals in freshwater and sea
water. . . .. . . 5

2 Heavy metal sulfide solubility products. 6

3 Relative advantages and disadvantages of
AAS, ASV, and ISE methods . . ... 10

4 Comparison of detection limits of AAS*,
ASV**, and ISE*** methods . . ... 11

5 Schedule for routine sampling. . ... 29

6 Aqueous heavy metal concentrations (pg/t)
at several sampling stations at the
Owens-Illinois site--October, 1975 to
August, 1976 . . . ... 50

7 Aqueous heavy metal concentrations (pg//)
in selected wells at the Owens-Illinois
site--October, 1975 to August, 1976. ... .52

8 Representative values of low, average,
and high concentrations of 0.1 N HNO3-
extractable metals in sediments from the
Owens-Illinois cypress domes.
(Concentrations in mg/kg dry weight) ..... 63

9 Distribution of metals between less mobile
and more mobile species in core S-1-0. ... .66

10 Heavy metal concentrations (mg/kg dry wt.)
in duckweed--April and May, 1976 . ... 79

11 Heavy metal concentrations (mg/kg dry wt.)
in cypress needles--May and June, 1976 . 81

12 Copper and cadmium concentrations (pg//)
associated with four molecular weight
fractions of natural and sewage organic 88


1









Table

13 Copper concentrations (pg/g wet wt.) in
duckweed cultured in two media at three
levels of other trace metals . .

14 Treatments used in the second duckweed
metal uptake study . . . .

15 Metal concentrations and total organic
carbon concentrations in solutions used
in the second duckweed metal uptake study .

16 Percent of total metal in duckweed released
after three weeks of decomposition . .


17 Ligand numbers and stability constants
for Cu and Cd complexes with the model
organic as determined by the Lingane
equation . . . *.

18 Effects of natural and sewage organic
fractions on DPASV peak potential. .

19 Effects of natural and sewage organic
on Cu and Cd ISE responses . ..

20 Copper complexation capacities and Ko
values for four model organic compounds

21 Copper complexation capacities and K
values for natural and sewage organics


* -113


* 126


* -127


. .131


. .141


vii


Page


. 93


. 95



. 95


.101















LIST OF FIGURES


Figure Page

1 Theoretical curves for determination of
complexation capacity of organic ligands
by the ion exchange equilibrium method of
Crosser and Allen (1976) . . .. 18

2 Site plan and groundwater monitoring wells
at the Owens-Illinois research site. ... .31

3 Sampling equipment and technique for shallow
wells. . . . ... .. .33

4 Percent water profiles in sediment cores
from the Owens-Illinois site . ... .58

5 Percent volatile solids profiles in
sediment cores from the Owens-Illinois site. 59

6 Percent sand profiles in sediment cores
from the Owens-Illinois site . .. .60

7 Percent clay and silt profiles in sediment
cores from the Owens-Illinois site . .. .61

8 Iron concentrations (mg/kg) in sediment
cores from the Owens-Illinois site . .. .67

9 Manganese concentrations (mg/kg) in
sediment cores from the Owens-Illinois site. 69

10 Copper concentrations (mg/kg) in sediment
cores from the Owens-Illinois site . .. .70

11 Mercury concentrations (mg/kg) in sediment
cores from the Owens-Illinois site . .. .72

12 Nickel concentrations (mg/kg) in sediment
from the Owens-Illinois site . ... .73

13 Lead concentrations (mg/kg) in sediment
from the Owens-Illinois site . .. .74


viii









Figure


Cadmium concentrations (mg/kg) in
sediment cores from the Owens-Illinois
site -


15 Zinc concentrations (mg/kg) in sediment
cores from the Owens-Illinois site .

16 Color and TOC profiles for organic
concentrates from the Austin Cary dome
and Newnans Lake eluted through Bio-Gel
P-2 . . . . .

17 Color and TOC profiles for organic
concentrates from Lake Mize and sewage
eluted through Bio-Gel P-2 . .

18 Color and TOC profiles for organic
concentrates from the Austin Cary dome
and sewage eluted through Bio-Gel P-6.

19 Molecular weight distributions of
organic concentrates as determined
by ultrafiltration . . .

20 Variations in copper uptake by duckweed
cultured in modified Hutner's medium .

21 Variations in copper uptake by duckweed
cultured in Hutner's medium less metals.

22 Copper and lead concentrations in
duckweed cultured in media receiving
various aliquots of natural and sewage
organic concentrates . . .

23 Cadmium and zinc concentrations in
duckweed cultured in media receiving
various aliquots of natural and sewage
organic concentrates . . .

24 Model organic compounds used in DPASV
and ISE studies. . . .

25 Effects of model organic on DPASV peak
current for 10-6 M . . .

26 Effects of salicylic acid, ferulic acid,
and phenylalanine on DPASV peak current
for 10-6M . . . .


.


Page


. 75


. 76




. 83



. 84



. 86



. 87


. 91


. 92




. 97




. 98


. .103


. .104



. .105









Figure


Effect of digallic acid on DPASV
peak current for 10-6 M . .

Effects of model organic on DPASV
peak potential for 10-6 M Cd . .

Effects of phenylalanine and digallic acid
on DPASV peak potential for 10- M Cu. .

Effects of salicylic and feruli? acids
on DPASV peak potential for 10 M Cu. .


. 106


. 110


S. 111


. 112


31 Effects of model organic on cadmium
ISE response . . . . .117

32 Effects of three model organic on
Copper ISE response . . .118

33 Effects of digallic acid on copper ISE
response . . . . .119

34 Effects of molecular weight fractions
of Austin Cary dome water on Cu and Cd
peak currents. . . . 122

35 Effects of molecular weight fractions
of sewage on Cd peak currents. . 124

36 Effects of molecular weight fractions
of sewage on Cu peak currents. . 125

37 Copper ion exchange isotherms for
Dowex 50X-8 and phenylalanine, salicylic
acid, and ferulic acid . . .. ... 132

38 Cadmium ion exchange isotherms for
Dowex 50X-8 and phenylalanine, salicylic
acid, and ferulic acid . . .. ... 133

39 Copper and cadmium ion exchange isotherms
for Dowex 50X-8 and digallic acid. . .134

40 Copper ion exchange isotherms for
Dowex 50X-8 and four molecular weight
fractions of Austin Cary dome water
concentrate . . . .137

41 Cadmium ion exchange isotherms for
Dowex 50X-8 and four molecular weight
fractions of Austin Cary dome water
concentrate. . . . . .138


Page









Figure


Page


42 Copper ion exchange isotherms for
Dowex 50X-8 and four molecular weight
fractions of sewage organic concentrate. .. 139

43 Cadmium ion exchange isotherms for
Dowex 50X-8 and four molecular weight
fractions of sewage organic concentrate. .. 140
















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



HEAVY METAL INTERACTIONS WITH NATURAL ORGANIC
IN AQUATIC ENVIRONMENTS

By

Neil Edward Carriker

June 1977

Chairman: Patrick L. Brezonik
Major Department: Environmental Engineering Sciences

Analyses of concentrations of eight heavy metals in

several components of cypress domes near Gainesville,

Florida which have received secondary sewage effluent for

three years showed that sediments and floating aquatic

plants are the principal heavy metal sinks in these

systems. Water samples from shallow wells surrounding the

domes showed no evidence of heavy metal contamination of

groundwater, although other parameters indicated that

contamination by sewage had occurred. Laboratory investi-

gations demonstrated the general utility of duckweed

(Lemna sp. and Spirodela sp.) in heavy metal phytotoxicity

studies and revealed that organic fractions from both

sewage and natural organic color complex significant amounts


xii








of heavy metals. Heavy metal interactions with these

natural and sewage organic and with four model organic

compounds were investigated by a variety of techniques

including ultrafiltration, atomic absorption spectropho-

tometry, ion-selective electrodes, differential pulse

anodic stripping voltammetry, and ion exchange equilibrium

methods; advantages and limitations of each technique are

discussed. Natural and sewage organic were observed to

have opposite molecular weight distributions. The various

analytical methods consistently indicated that both the

natural and sewage organic and model compounds completed

Cu much more strongly than Cd, and indicated that the

complexation capacities of the natural and sewage organic

increased with increasing molecular weight.


xiii















CHAPTER I
INTRODUCTION

Research on heavy metal cycling in the environment has

received increased attention and popularity in recent years

as a result of public concern about pollution of lakes and

streams by sewage and the subsequent passage of the Federal

Water Pollution Control Act of 1972, which requires "zero

discharge of pollutants" by 1985. The enormous costs of

achieving "zero discharge" by conventional waste-treatment

methods and the stresses our burgeoning population has

placed on our current freshwater resources have created con-

siderable interest in utilizing land disposal of sewage ef-

fluents as a low-cost means of achieving tertiary treatment

of waste water and simultaneous recharge of freshwater

aquifers.

The toxicities of heavy metals and our limited under-

standing of their interactions with natural organic raise

questions of potentially serious public health risks from

activities which involve disposal of heavy metal-contaminated

wastes such as sewage, sewage sludge, and industrial and

mining wastes. One of the principal concerns is that

solubilization of heavy metals, partly through complexation

reactions with organic ligands, may promote heavy metal

transport through the soil profile resulting in groundwater

contamination. The effect of complexation on heavy metal

1









toxicity is equally important. The majority of investiga-

tions support the position that complexation reduces

phytotoxicity by reducing the concentration of free metal

ions in solution; however, some investigators have reported

increased metal uptake in the presence of certain organic

ligands. Two factors which may exert controlling influ-

ences on heavy metal phytotoxicities are the kinetic

reactivities liabilities ) of heavy metal-ligand complexes

and possible synergisms between different metals.

The major impediments to better understanding of heavy

metal cycling in the environment have been poor understand-

ing of the nature of natural organic present in aquatic

ecosystems and a lack of analytical methods yielding un-

equivocal results about heavy metal speciation in environ-

mental samples. Thus, despite numerous investigations in

recent years, there are still far more questions than

answers about the nature and extent of heavy metal-organic

interactions in the environment.

The research reported here is part of a larger inter-

disciplinary research effort at the University of Florida

dealing with recycling wastewater through cypress wetlands

ecosystems. In this project several small cypress domes

near Gainesville, FL, are receiving secondary sewage

effluent from a trailer park. The responses of the swamp

ecosystems to the sewage effluent and the effectiveness of

the swamps in removing pollutants from the effluent are

being monitored to determine the feasibility of using








cypress wetlands as natural tertiary treatment systems for

municipal waste effluents.

The heavy metals phase of this larger research effort

has been aimed at clarifying the following questions:

1) What are the ultimate fates of heavy metals in

sewage added to the domes?

2) What is the potential for groundwater contamina-

tion from these activities?

3) What is the nature of the interactions between

heavy metals and naturally-occurring organic

compounds and organic present in sewage?

4) What role do these compounds play in detoxifica-

tion of heavy metals and in promoting heavy metal

transport through soils?

Obviously, ultimate resolution of these questions will

require much more research than could be presented in one

doctoral dissertation. Indeed, the exact answers to these

questions will vary from one location to another as vari-

ables such as soil type, climate, and relative concentra-

tions and types of metals and organic change. The results

reported here will further our understanding of the

environmental chemistry of heavy metals and will help

determine the feasibility of using wetlands ecosystems for

sewage recycling.















CHAPTER II
LITERATURE REVIEW

Several reviews (Singer 1973; Rubin 1974; Klein 1976)

recently have thoroughly considered various aspects of heavy

metals in the environment. Another exhaustive review of

the same literature is neither necessary nor desirable;

therefore the emphasis of this chapter is twofold: (1) to

create some perspective on the problems of heavy metals in

the environment by briefly summarizing the most important

points covered in existing literature reviews, and (2) to

supplement the existing reviews by considering work pub-

lished in the current year. Four areas pertinent to the

work reported here will be considered in the following

discussion. These are (1) general considerations in the

environmental chemistry of heavy metals, (2) analytical

methods suitable for heavy metal analysis of environmental

samples, (3) investigations of heavy metals in sewage and

sewage sludge, and (4) investigations of heavy metal-

organic interactions in aquatic ecosystems.

A. Environmental Chemistry of Heavy Metals

Recent papers which deal with the mechanisms and para-

meters controlling heavy metal solubilities in the environ-

ment include chapters by Andelman (1973), Morel et al. (1973),

Leckie and James (1973), and Williams et al. (1974) in the

two books edited by Singer (1973) and Rubin (1974). Jenne

4









and Luoma (1975) presented a review emphasizing the biologi-

cal implications of processes which promote solubilization

of heavy metals, Eisenreich et al. (1976) considered factors

contributing to dissolution of metal sulfide minerals, and

Klein (1976) prepared an exhaustive review of the environ-

mental chemistry of the eight heavy metals studied in this

research. Table 1 summarizes typical concentrations of

these metals observed in freshwater and sea water by various

workers.


Table 1. Typical concentrations of selected heavy
metals in freshwater and sea water

Concentration, pg/

Metal Freshwater Sea Water References

Fe ~30* 10 Kopp & Kroner (1967);
Williams et al. (1974)
Mn 50* 2 Kopp & Kroner (1967);
Horne (1969)
Zn 10-50 10 U.S.G.S. (1971);
Horne (1969)
Cu 15 3 Kopp & Kroner (1967);
Horne (1969)
Cd <1 -0.15 Bowen (1966)
Pb 1-50 0.03 U.S.G.S. (1971);
Horne (1969)
Ni 19 2 Kopp & Kroner (1967);
Goldberg (1965)
Hg 0.02-0.7 0.02-0.7 D'Itri (1972)

*Much higher concentrations (one to two orders of magnitude)
are found in anoxic waters.


Three types of mechanisms dominate the phase equilibria

of heavy metals in environmental systems: (1) precipitation

reactions with inorganic ligands, principally hydroxide,









carbonate, and sulfide, (2) complexation reactions with

inorganic and organic ligands, and (3) sorption and ion

exchange reactions with colloidal particulate matter and

clays.

Sulfide solubility equilibria are dominant for most

heavy metals under reducing conditions and pH values greater

than 7. Leckie and James (1974) have listed the principal

species formed and their solubility products (Table 2).

The low Ksp values illustrate the extremely low solubili-

ties of most of these species.


Table 2. Heavy metal sulfide solubility products


Species PKsp @ 250C

FeS 18.8
FeS2 (pyrite) 26.1
MnS 15.7
ZnS 22.0
CuS 36.2
CdS 27.2
PbS 28.2
a-NiS 20.8
Y-NiS 27.8
HgS (cinnabar) 53.6
Reference: Leckie and James (1974)


Fe and Mn are exceptions to the general rule of reduced

solubilities of heavy metals under anoxic conditions. These

two metals exist in two valence states in natural waters,

with the reduced form (+2 valences) predominant in anoxic

waters. The amount of Fe available usually exceeds the

amount of sulfide present, and the amount of Mn frequently

exceeds the amount of sulfide. Under these conditions









their solubilities are governed by carbonate equilibria,

and their solubilities of ferrous and manganous carbonates

are higher than the solubilities of the oxide, hydroxide,

phosphate and carbonate salts of the oxidized metals which

dominate the aerobic phase equilibria of Fe and Mn.

Carbonate equilibria generally control heavy metal

solubilities in aerobic natural waters up to pH values of

about 8 to 10, above which hydroxide and oxide equilibria

begin to dominate the system. Mixed carbonate, hydroxide

salts (e.g., apatite, malachite, etc.) also occupy a

prominent role in heavy metal solubility equilibria in

moderately alkaline waters. Free metal ions generally are

the predominant species in solution at pH values less than

6 if no organic ligands are present. The principal in-

soluble Hg species in aerobic waters are HgC12 and Hg2C12

(pKsp = 13.8 and 18.0) at pH < 7 and HgO (pKsp = 25.7) at

pH > 6 (Gavis and Ferguson 1972).

Several attempts to determine the relative stabilities

of complexes of heavy metals with natural organic matter

have produced conflicting results, depending on the condi-

tions of investigation, sources of organic matter, and in-

vestigative techniques (Irving and Williams 1948; Schnitzer

and Skinner 1967; Schnitzer and Hanson 1970). While these

results indicate that serious problems exist in attempts to

develop exact sequences of relative stabilities, they also

indicate that under the conditions present in natural

waters the following general trends are apparent: Cu and


1









Fe form the strongest complexes, Mn and Zn form the weakest

complexes, and Pb, Ni, and Cd occupy intermediate positions.

Mercury also forms relatively strong organic complexes

(Miller 1975; Lindberg and Harriss 1974); in addition, it

forms a variety of organo-mercurials in which mercury is

covalently bonded to carbon.

Sorption and ion exchange processes also occupy a

prominent role in heavy metal environmental chemistry.

Sorption equilibria between free metal ions in solution and

metals sorbed onto colloidal particulate matter and sedi-

ments are important mechanisms buffering the effects of

"shock" loadings of heavy metals to aquatic systems. Leckie

and James (1974) state that for each hydrolyzable metal ion

species "there is a critical pH range, often less than 1

unit wide, over which the fractional amount of metal ad-

sorbed increased from almost zero to unity," with different

substrates usually having only small effects on the critical

pH range. The presence of organic ligands may either reduce

the amount of sorption by forming soluble non-sorbing com-

plexes, or it may increase sorption by forming readily

sorbed complexes.

In addition to the reviews by Jenne and Luoma (1975)

and Eisenreich et al. (1976), two papers not covered in

previous reviews should be noted. Hahne and Kroontje (1973)

calculated the species distributions of hydroxide and chlor-

ide complexes of Hg, Cd, Zn, and Pb under varying conditions

and concluded that these complexes may be important factors








in the movement of heavy metals in soil solutions and down-

stream reaches of some rivers. Sylva (1976) used a compu-

terized equilibrium model to investigate the aqueous envi-

ronmental chemistry of copper. He concluded that (1) in the

absence of organic ligands precipitation of malachite,

Cu2(OH)2CO3, is the thermodynamically favored process for

removal of free Cu (II) from freshwater, (2) that adsorp-

tion of free Cu (II) by sand, gravel, and suspended clay

is important at pH > 6.5, and (3) that organic ligands which

form copper complexes with stability constant, 8, less than

4 exert a negligible influence on these processes at pH 6

and pH 7, while ligands with 8 > 5 are the principal agents

for reduction of free Cu concentrations at these pH values.

B. Analytical Methods Applicable to Heavy
Metal-Organic Interaction Studies

1. Instrumental Methods for Measuring Heavy Metal
Concentrations

Several instrumental methods of analysis, including

atomic fluorescence spectroscopy, spark and x-ray emission

spectroscopy, and neutron activation analysis, are suitable

for measuring heavy metal concentrations at the pg/t levels

normally encountered in the natural aquatic environment;

however, atomic absorption spectrophotometry (AAS), po-

larography and anodic stripping voltammetry (ASV) and ion-

selective electrode (ISE) techniques have been the methods

of choice among most environmental scientists. Relative

advantages and disadvantages of these three methods are

presented in Table 3, and detection limits are listed in

Table 4.














Relative advantages and disadvantages
of AAS, ASV, and ISE methods


Method


Advantages

1. Speed of Analysis
2. Applicable to a large
number of elements
3. Detection limits in
the pg/1 range
4. High selectivity for
a single element--
relative freedom
from interference
from other elements




1. Simultaneous analy-
sis of several metals
2. Discrimination be-
tween completed and
uncomplexed forms
3. Detection limits in
the sub-pg/l range
4. Large dynamic range

1. Simplicity of method
and instrumentation
2. Speed of Analysis
3. Measures only
activity of free
metal ion in solution


Table 3.


AAS













ASV








ISE


Disadvantages

1. Measures total metal
present--incapable of
discriminating between
species
2. Non-atomic absorption
interference
3. Matrix effects
4. Formation of refractory
compounds and inter-
metallic compounds
5. Ionization suppression
or enhancement by
alkali metals

1. Limited to almalgam-
forming metals
2. Time-consuming
3. Synergistic effects of
Cu and Zn
4. Interferences due to
sorption effects


1. Interferences from
other metals limit
usefulness for environ-
mental samples
2. Detection limits ~10 M
3. Electrodes available
only for Ca, Mg, K, Na,
Ag, Cd, Cu, and Pb
4. Sensitivity limited by
solubility of metal
sulfides used in elec-
trode construction,
and by instability of
reference electrodes.









Table 4. Comparison of detection limits
of AAS*, ASV**, and ISE*** methods


Detection Limit, pg/

AAS ASV

non- linear differential
Metal flame flame scan pulse ISE

Fe 5.0 0.8 --- --- --
Mn 3.0 0.1 --- -- --
Zn 2.0 0.02 0.04 0.04 --
Cu 3.0 0.3 0.01 0.005 0.6
Cd 0.6 0.02 0.01 0.005 10
Pb 15 0.2 0.02 0.01 20
Ni 8.0 1.0 --- --- --
Hg 200 0.04 -- --


* Parker (1972)
** Copeland and Skogerboe (1974)
*** Orion Research, Inc. (1971, 1968, 1972)



Since the introduction of the first commercially avail-

able atomic absorption spectrophotometers about a decade

ago, AAS has become the workhorse of trace metals analysis.

During analysis, a line spectrum emitted from a lamp in

which the cathode contains the element of interest is passed

through an area in which ions present in a sample are re-

duced to the atomic state. Atoms of the element being de-

termined absorb the incident radiation in proportion to

their concentration in the sample. Normally, atomization

is effected by direct aspiration of sample into a flame,

but recent improvements utilizing electrically heated

graphite atomizers, which greatly improve detection limits,

are becoming more common. The method is relatively








sensitive, versatile, and simple: about 60 elements can be

analyzed at the mg/g level and detection limits approach

the pg/t level for about 20 of these. Spectral interfer-

ences from other elements are rare; the principal interfer-

ences are absorption and light scattering by non-atomic

species in the light path, matrix effects which affect the

efficiency of atomization, and chemical interference such

as formation of refractory oxides, phosphates, and silicates

and suppression or enhancement of ionization by alkali met-

als in the samples. One major deficiency of the method for

environmental analysis is its inability to discriminate be-

tween different forms of the element present in a sample.

ASV is a two-step electroanalytical method: metal

ions in solution are first reduced and plated into an elec-

trode held at a sufficiently negative potential, then the

electrode potential is increased linearly with time, caus-

ing the metals which have been concentrated in the electrode

to be oxidized back into the solution. The diagnostic para-

meters are peak potential, E and peak current, i E is

a function of the metal involved, the reversibility of the

reaction, and the concentration and type of ligand present,

and i is a function of metal concentration, deposition

time and voltage, stirring rate, voltage scan rate anodicc

stripping step), electrode surface area, and concentration

and type of ligand present. If all other variables are

held constant i is directly proportional to the concentra-

tion of metal in solution. The principal variations in the








technique are alternating current stripping voltammetry

(ACSV), in which a small amplitude ac signal is superim-

posed on the linear ramp potential, and differential pulse

anodic stripping voltammetry (DPASV), in which a larger

amplitude (25-100 my), regularly spaced pulse waveform is

superimposed on the ramp. Both variations utilize sophis-

ticated signal analysis techniques to reduce the nonfaradic

(capacitative) component of the current, thereby increasing

sensitivity (Copeland and Skogerbee 1974; Flato 1972).

The principal advantage of ASV is its applicability to

metal speciation studies. Several investigators (Allen

et al.1970; Mancy 1972; Ernst et al.1975; O'Shea and

Mancy 1976; Gachter et al. 1973; Chau et al. 1974; Chau and

Lum-Shue-Chan 1974) have reported variations of procedures

utilizing titration of a ligand system with a metal titrant,

or vice versa, and frequently including analysis before

and after mild acid hydrolysis, to draw conclusions about

the completing capacities of natural waters (see section 3).

These procedures are based on the assumptions that only free

metal ions are reduced at the electrode, that only non-

reducible complexes are formed, and that acid hydrolysis

destroys all complexes formed. These may be valid assump-

tions in many cases; however, Brezonik et al. (1976) have

demonstrated that interpretation of the results may not be

as straightforward as has been suggested, especially in view

of the currently limited understanding of the types of

complexes formed with natural organic.








The only heavy metal ion-selective electrodes commer-

cially available are for Cu, Cd, Pb, and Ag. These elec-

trodes are potentiometric devices which measure the potential

developed at the surface of a metal sulfide crystal by the

activity of that metal ion in solution. The Nernst equation

describes the potential/activity relationship:


E = E + .0591 log ai (1)


where a. = metal ion activity in solution, E = standard

electrode potential for the metal ion being measured, and

n = number of electrons involved in the redox reaction.

ISE's respond only to free metal ion activity and not to

any completed forms. Their use in routine monitoring of

natural waters is generally limited by high detection limits

relative to concentrations in natural waters and chemical

interference; their unique characteristics are better

utilized in laboratory investigations at slightly higher

concentrations under controlled conditions.

2. Methods for Determining Molecular Weight Distributions
of Organics

Gel permeation chromatography (GPC) is a technique in

which compounds are separated by molecular weight on the

basis of the routes they follow in being eluted through a

column containing porous gel beads manufactured from syn-

thetic cross-linked polymers. Small, low molecular weight

compounds penetrate the pores of the gel and are eluted

from the column more slowly than large molecules which








travel only in the void volume outside the gel. V the

excluded (external, void) volume of the column is the

volume at which the largest molecules are eluted, and Vt,

the total column volume, is the volume at which small ionic

species are eluted. Vi, the included volume of the gel is

the difference, Vt-Ve. A variety of gels are commercially

available with different ranges of pore sizes and surface

characteristics. Klein (1976) has described the method in

greater detail and has reviewed GPC applications in water

chemistry. He notes that a major difficulty with the

method in heavy metal-organic interaction studies is that

the gels are not chemically inert; sorption of metal ions

by charged sites is a serious problem with certain gels and

low ionic strength eluants.

Ultrafiltration techniques separate compounds into

various molecular weight ranges on the basis of their

ability to pass through a series of membrane filters with

nominal pore sizes ranging from 2.4 to 12 nm. The low

membrane surface/solution volume ratios characteristic of

this method and the non-ionic nature of the membranes mini-

mize sorption problems. The major caution to observe with

ultrafiltration is to leach all soluble organic material out

of new membranes by rinsing with copious volumes of dis-

tilled water prior to use (Smith 1976).

3. Methods for Observing Heavy Metal-Organic Interactions

Several investigations of heavy metal-organic inter-

actions in natural waters have depended on the ability of








ASV to distinguish between electrochemically labile and non-

labile forms of metals in solution, i.e., metal species

which are reducible at the mercury electrode (labile), and

species which are not reducible at the electrode (non-

labile). These studies have included titration of natural

water samples with aliquots of standard metal solutions

(Chau et al. 1974; Mancy 1972), titration of solutions of

metal ions with ligand concentrates (Ernst et al. 1975), or

a combination of both procedures (O'Shea and Mancy 1976).

Additional ASV procedures have included analysis before

and after mild (Allen et al.1970) or rigorous (Chau and

Lum-Shue-Chan 1974) acid hydrolysis, and variation of the

voltage scan rate in the anodic stripping step (Mancy 1972).

Refinements of these techniques have coupled ASV analysis

with GPC (Bender et al. 1970) or ultrafiltration (Smith

1976) fractionation to determine the molecular weight range

of the active completing agents.

Schindler and Alberts (1974) combined ultrafiltration

with AAS in a study of metal-organic interactions in four

Georgia lakes, and Lam (1976) used an ISE to monitor free
2+
Cu2 activity in titrations of soil fulvic acid with copper

sulfate solution.

Mantoura and Riley (1975) described a GPC/AAS tech-

nique which apparently reduces the sorption problems noted

earlier. In their procedure a gel column is equilibrated

with a buffer solution containing a fairly high concentra-

tion of free metal (~10 mg/i). An organic concentrate









which has been equilibrated with the buffer solution is then

eluted through the column, and the collected fractions are

analyzed by AAS. A plot of metal concentration versus vol-

ume of eluant shows a peak due to completed metal at the

same volume in which the organic ligand is eluted, and a

negative peak (metal deficiency) appears at the volume

corresponding to elution of low molecular weight solutes.

The areas of the two peaks determine the amount of metal

completed, and knowledge of the initial ligand concentration

and the background metal concentration permits calculation

of stability constants. Multiple chromatographic runs at

varying metal concentrations can yield information about

different metal binding sites on one ligand or sequential

complexation by a mixture of ligands.

Crosser and Allen (1976) have developed an ion ex-

change equilibrium method for determination of complexation

capacities of soluble organic ligands based on earlier work

by Zunino (1972). The method involves competition between

an organic ligand or ligand mixture and a strong cation

exchange resin. A controlled excess of ion exchange resin

and varying amounts of a metal stock solution are added to

a series of vessels, each containing an equal volume and

concentration of ligand solution. The systems are equi-

librated by shaking 24 hours, and the supernatants are

analyzed for total soluble metal by AAS. Plots of metal

bound by the resin versus total soluble metal are of three

possible forms (Figure 1), depending on the stabilities of




















z
U;
QZ



-J
W













0
im
_j
W


I
TOTAL METAL CONCENTRATION
IN SOLUTION COMPLETEDD & FREE)


Figure 1.


Theoretical curves for determi-
nation of complexation capacity
of organic ligands by the ion
exchange equilibrium method of
Crosser and Allen (1976). (A:
no ligand present; B: strong
ligand; C: mixture of two
ligands, both weaker than the
ligand in B.)








the metal-ligand complexes formed, the ligand concentra-

tion, and the number of ligands present. The principal

advantages of this method are its applicability to mixed

ligand systems and its avoidance of potential errors as-

sociated with speciation techniques since it requires only

total soluble metal analysis.

C. Investigation of Heavy Metals in Sewage

Most investigations of heavy metals in sewage and sew-

age sludge have been concerned with possible contamination

of agricultural soils and crops by use of sewage effluent

for irrigation and use of sludge as a soil conditioner.

Jorgensen (1975) has reviewed several of these investiga-

tions, and studies by Bradford et al. (1975) and Kirkahm

(1975) dealing with soil analysis and crop uptake indicate

the general nature of the problem and the investigative

techniques.

Model ecosystem studies by Lu et al.(1975) demonstrated

a clear cut mobilization and transfer of Cu, Cd, Pb, and

Zn into the food chains and biota after application of sew-

age sludge, with Cd exerting a particularly adverse effect.

Mobilization of these metals was most pronounced in sandy

soils with low cation exchange capacities (CEC), and was

lowest in silty clay loam soils with high CEC's.

Lund et al. (1976) extracted Cu, Cd, Ni, Zn, and Cr fron

cores taken from sewage disposal ponds and found elevated

levels in surficial sediments. High correlations of metal

profiles with chemical oxygen demand profiles suggested








that soluble metal-organic complexes may be an important

metal transport mechanism in these soils.

McDonald and Clesceri (1971) utilized GPC techniques

to examine the effects of wastewater organic fractions on

the growth of Selenastrum capricornutum and Anabaena flos-

aquae. Bender et al. (1970) demonstrated the complexation

of heavy metals by sewage organic, and Culp (1975),using a

combined GPC/DPASV approach, obtained results which sug-

gested that the copper-binding ability of sewage organic

depended on the biodegradability of the organic matter

present.

Klein (1976) has reported results for the first two

years of investigations on heavy metal cycling in cypress

wetlands ecosystems receiving secondary sewage effluent;

this dissertation presents results for the third year of

that study. Klein found that heavy metal levels in stand-

ing waters of cypress domes receiving domestic sewage were

slightly higher than levels in control domes, but levels in

both standing waters and shallow wells were within the range

of typical concentrations in natural waters (see Table 1).

Slightly elevated concentrations in shallow groundwater

downstream from the sewage domes suggested that some metal

loss occurs through the domes' floors. Laboratory studies

showed that cadmium and zinc uptake by duckweed was unaffected

by the presence of sewage organic or natural organic color,

but copper uptake was reduced in both cases. DPASV and AAS

analyses of GPC fractions indicated that Cd, Mn, and Zn








exist as free ionic species or in association with low

molecular weight organic, and Cu exists in association

with both low and high molecular weight compounds.

D. Investigations of Heavy Metal-Natural
Organic Interactions

Numerous investigations and several symposia in the

last decade have focused on the role of natural and anthro-

pogenic organic compounds in the biogeochemistry of heavy

metals (Faust and Hunter 1971; Hood 1970; USERDA 1975).

Although experimental conditions and methods have varied

considerably between investigations, the results have usu-

ally been consistent with the following general observations

about the nature of naturally-occurring organic compounds in

aquatic environments and their interactions with heavy metals.

Concentrations of dissolved organic compounds in natural

Sweaters normally range from 0.1 to 10 mg/t, but values as high

as 30 to 40 mg/t may be observed in highly colored bogs and

lakes. Organic color, dissolved organic matter (DOM), natu-

ral organic matter (OM), yellow organic acids, fulvic acids

(FA), humus, and humic acids (HA) are the most common terms

used to describe this class of compounds. Soil scientists

distinguish between humic and fulvic acids on the basis of

molecular weight and solubility in dilute mineral acid and

dilute base, but these and the other terms have been used

somewhat interchangeably in the aquatic chemistry literature.

There has been some controversy about whether organic

color compounds are predominately aliphatic (Shapiro 1957),

cyclic aliphatic (Ishwatari 1973), or aromatic (Black and








Christman 1963), but there is little doubt that they are

primarily composed of carbon, hydrogen, and oxygen, with

perhaps 1 to 2 percent nitrogen present. GPC and ultra-

filtration studies have indicated that most natural organic

fall in the 1,000 to 100,000 Dalton molecular weight range,

with the specific distribution varying with source and his-

tory (Gjessing 1971; Black and Christman 1963; Christman

and Ghassemi 1966; Schindler and Alberts 1974).

As the terms used to describe them suggest, natural

color organic exhibit acidic properties. This acidity is

due to a preponderance of carboxylic acid, phenolic and

alcoholic functional groups on organic color macromolecules,

which are also responsible for the observed complexation

capacity of these compounds. Christman and Ghassemi (1966),

Stevenson and Ardakani (1972), and Gamble and Schnitzer

(1973) have proposed hypothetical structures for natural

organic compounds which account for their elemental com-

position, degradation products, acidic properties, and

complexation capacities. These structures include adjacent

carboxylic acid and phenolic groups as one possible

mechanism for chelation of metal ions (Gamble and Schnitzer

1973):

0 O
II II
0 ~ + CU ++ +

OH 0 "H
0/C








The reported relative stabilities of heavy metal-

organic complexes apparently vary slightly depending on

the conditions of investigation, but the highest stability

constants have consistently been reported for copper com-

plexes (see section A). This may account for the reduced

phytotoxicity of Cu in highly organic sulfate liquor wastes

from pulp mills (Wilson 1972), and for decreased algal

growth in sea water in which dissolved organic had been

photo-oxidized with UV light (Barber 1973). Complexation

of other heavy metals may also reduce their phytoxicities,

since it is generally believed that toxic effects usually

are due to uptake of free metal ions which inhibit essen-

tial enzyme reactions. Increased solubility of heavy metals

by complexation with natural organic may also promote pri-

mary productivity in some cases by providing a reservoir

of non-toxic micronutrients in solution which may be reduced

to forms available for biological uptake at the plant-

solution interface. Martin and Martin (1973) have proposed

that solubilization of iron by humic substances may be an

important factor in stimulating blooms of red tide organisms

in the Gulf of Mexico, but an alternate explanation may be

reduction of copper toxicity by complexation. However, in-

creased solubility due to complexation may also facilitate

heavy metal transport through soils into groundwater.

Mantoura and Riley (1975) found evidence for two dif-

ferent types of metal binding sites on fulvic acids ex-

tracted from lake water, peat, and soil. Apparently the








same sites were involved in complexation of Cu, Ni, and Zn,

but fulvic acids from peat contained about 1.5 times more

sites per unit weight and formed complexes with stability

constants about half an order of magnitude greater than

fulvic acids from lake water. Stability complexes for Cu

complexes with peat and lake water organic were about three

orders of magnitude greater than for the corresponding Ni

and Zn complexes, but little difference was observed be-

tween stabilities of Cu, Ni, and Zn complexes with soil

fulvic acid.

Schindler and Alberts (1974) observed significant dif-

ferences in size and composition of organic compounds in

aerobic and anoxic waters in four Georgia lakes, and Lam

(1976) observed significant differences in Cu completing

abilities of organic matter extracts from Florida and Viet-

namese soils. O'Shea and Mancy (1976) observed formation

of strong, non-labile complexes between Cu and commercially

available humic acid, while Tl formed predominately labile

complexes, and Cd exhibited intermediate behavior.

Benes et al. (1976) found evidence for extensive com-

plexation of Hg and Zn by humus in a Norwegian stream, with

Fe associated with colloids or complexes with molecular

weight greater than 10,000 Daltons. Miller (1975) also

observed strong complexation of Hg by humic acids in fresh-

water sediments.

Hutchinson and Czyrska (1975) observed synergistic ef-

fects of Ni on Cd uptake by duckweed, and Stokes (1975)









observed similar synergisms in investigations of Cu and Ni

uptake by metal-tolerant strains of Scenedesmus.

In summary, complexation by natural and anthropogenic

organic compounds is generally regarded as an important

mechanism for detoxification and transport of heavy metals

in the environment, and may significantly affect the bio-

geochemical cycling of these elements. Numerous investiga-

tions have used a variety of analytical techniques to ex-

amine the nature of organic color compounds and their inter-

actions with heavy metals, but the complexity of the mixture

of compounds which constitute humicc substances," variations

between sources, and limitations of analytical methods have

hampered efforts to determine the active sites of complexa-

tion and relative stabilities of different metal-organic

complexes.

"The major difficulty in determination of
speciation of trace elements among the
various solute forms is the lack of ade-
quate techniques to estimate the quantity
and stability of the complexes formed
with natural organic ligands chemi-
cal assessment of availability to aquatic
fauna as a function of physiochemical
form and determination of the quantity of
and stability constants for natural trace
element-organic complexes" are among the
most pressing research needs
(Jenne and Luoma 1975).















CHAPTER III
AREAS OF INVESTIGATION, SAMPLING PROCEDURES,
AND ANALYTICAL METHODS

A. Field Studies

Field investigations were conducted in three areas to

further characterize heavy metal cycling in cypress domes.

Routine monitoring of heavy metal levels in standing wa-

ters, shallow wells, and sewage, begun by Klein (1976), was

continued, and characterization of metal profiles in sedi-

ments and investigations of heavy metal levels in duckweed

(Lemna sp. and Spirodela sp.) and cypress needles (Taxodium

sp.), also begun by Klein, were continued and expanded.

1. Site Description

The main study site consists of three cypress domes

located on a large pine plantation about five miles north-

west of Gainesville, FL, and owned by Owens-Illinois, Inc.

Two of the three domes, control dome I (C-l, 0.69 ha), and

sewage dome I (S-l, 0.51 ha), were extensively burned in a

forest fire December 4, 1973, while the third dome, sewage

dome II (S-2, 0.99 ha) was burned along the southwest edge.

A "package treatment plant" serving a 155-unit mobile

home park adjacent to the Owens-Illinois site supplies

secondarily treated sewage effluent to S-1 and S-2 at a

current rate of 2.5 cm/wk. The sewage plant is an extended









aeration waste treatment facility which is currently treat-

ing about 95 m3/day (25,000 GPD), 83 percent of its design

capacity of 114 m3/day (30,000 GPD).

Sewage application rates to S-1 have varied somewhat

during the course of the project. Pumping was initiated in

March, 1974, and continued at rates varying from 7.6 to

12.7 cm/wk until November, 1974. Pumping was then halted

from November, 1974, to March, 1975, to allow the dome to

dry out and consolidate the gelatinous muck which had built

up, and was resumed in March, 1975, at the rate of 2.5 cm/wk

and continued until July, 1976.

A combination of high sewage application rates, heavy

rainfall, and abnormally high surface runoff during summer,

1974, caused overflow from S-1. A weir was constructed on

the west side of the dome to maintain the standing water at

a normal level and to allow the overflow to be measured and

sampled.

S-2 functioned as a control during most of 1974, but

has been receiving sewage at the rate of 2.5 cm/wk since

November, 1974. C-l has received groundwater from a deep

well (-50 m) according to the same schedule and loading

rate as S-1 to separate the response of acidic cypress

domes to application of hard, alkaline groundwater from the

effects of sewage pollutants (nutrients, heavy metals, or-

ganic matter).

One larger dome (-4.5 ha), located in the Austin Cary

Forest about four miles east of the Owens-Illinois site,








serves as a "natural" control, i.e., no additional water is

pumped into this dome, although other activities--well drill-

ing, installation of a boardwalk, tower, equipment shed,

and various monitoring devices--have somewhat altered the

natural character of this dome.

2. Routine Monitoring

The standing waters, sewage, and shallow groundwaters

on the Owens-Illinois site have been routinely monitored

for heavy metals, major cations, and other water quality

parameters since March, 1974. This dissertation deals pri-

marily with the routine monitoring effort since September,

1975. Results from the March, 1974, to September, 1975,

period are briefly summarized in Chapter II, but detailed

sampling schedules, procedures, and results for this period

have been described by Klein (1976).

Samples were collected on a monthly basis from October,

1975, to February, 1976, and on a quarterly basis in May

and August, 1976, from the stations listed in Table 5.

Well 29 at C-l, wells 6A and 21 at S-1, and wells B-2, B-4,

B-6, B-10, B-ll, and B-12 at S-2 (see Figure 2.) were also

sampled during the period October to February. These and

the wells listed in Table 5 are shallow (10-20 ft. depth)

and just penetrate a clay hardpan stratum, which is a

prominent feature of the area's stratigraphy.

The procedure for sampling wells involved pumping each

well dry (or removing at least 4 liters) two to three days

before sampling so fresh percolate could be collected.



















Table 5. Schedule for routine sampling


Site


# Samples

2


Sewage


Groundwater
Well 33


A.C.


C-l


S-1


S-2


Descriptions

Treated effluent, oxidation
pond effluent
Deep well supplying C-l
Background for water table
aquifer
Standing water, Austin Cary
dome
Standing water: center,
30 m west
Wells 28 and 30
Standing water: center, 15 m,
and 30 m east
Wells 4A, 8A, 19, and 22
Weir overflow
Standing water: center, 40 m
north, 30 m south
Wells B-3, B-5, B-7, and
B-9


































Figure 2. Site plan and groundwater monitoring wells

at the Owens-Illinois research site.



























































OWENS-ILLINOIS
RESEARCH SITE
CENTER FOR WETLANDS
UNIVERSITY OF FLORIDA

GROUNDWATER MONITORING WELLS


GROUNDWATER
DOME I


Point of
Groundwater
Discharge








Samples were collected with a hand vacuum pump (Nalgene) by

drawing water through 1/4 in i.d. Tygon tubing that was

custom-fitted to each well (down to the perforated zone,

1-2 ft above the bottom of the casing) (Figure 3). The

first liter of water drawn was discarded to clean the tub-

ing of contaminants. Samples were then collected directly

in acid-washed 125-ml polyethylene bottles. All other

samples were "grab" samples.

Within three hours of collection, all samples were trans-

ported to the laboratory, where they were filtered through

acid-washed glass fiber filters (Gelman Type GF/A) to re-

move particulate matter, acidified with 1 ml redistilled

HNO3, and frozen until analyzed.

Samples were analyzed for Fe, Mn, Zn, Cu, Cd, Pb, Ni,

and Hg by atomic absorption spectroscopy (AAS). In addition,

all samples except those collected in December, 1975, and

January, 1976, were analyzed for major cations (Ca, Mg, Na,

K). All analyses were performed using a Varian Techtron

Model 1200 Atomic Absorption Spectrophotometer with the in-

strument settings recommended by the manufacturer. Sample

atomization was by an air/acetylene flame (major cations,

Fe, Mn, Zn, Ni), a graphite tube furnace (Varian Techtron

Model CR-63; Cu, Cd, Pb), or an apparatus (Varian Hg/As/Se

Analysis Kit Model 64) utilizing wet chemical oxidation
2+
(5% H2SO4, 0.1% KM 04) of organic mercurials to Hg fol-

lowed by reduction by stannous chloride to elemental mer-

cury, and purging with air to strip volatile mercury vapors



















Well Casing-

1/4" Sample Tubing-


Pump


Sterile
Sample
Bottle


Level


Figure 3. Sampling equipment and technique for
shallow wells.








into a quartz absorption tube (Hg). Since nickel levels

were near the limits of detection by direct sample aspira-

tion into an air/acetylene flame, and since the graphite

tubes used with the CR-63 atomizer were found to contain

appreciable Ni contamination, Ni values reported here were

determined using the INTEGRATE-10 mode built into the

Model 1200. In this mode, the output signal is integrated

for a ten second interval before being displayed on the

readout. This enhances the instrument's sensitivity by av-

eraging the random fluctuations due to background noise,

enabling better detection of low-level signals. The Cu,

Cd, and Pb analyses for the August, 1976, samples were also

performed in this manner.

3. Sediment Characterization and Metal Profiles

A general characterization of the sediments in C-l,

S-l, and S-2, and a determination of metal profiles in these

sediments was conducted in the following manner.

Sediment cores were obtained in domes S-l, S-2, and

C-l by attaching a sharpened coring tip to a section of

1-1/2" O.D. aluminum tubing and driving it into the sediments

with a sledge hammer to a depth of 25-50 cm. The tubing was

then capped and withdrawn and the core inside was extruded

into plastic bags in the field, sectioning it into 3-cm or

5-cm sections. Two cores were obtained in this manner at

the center of S-2 in November, 1975, and eight additional

cores were obtained in March, 1976, at locations 0, 10, and

20 meters from the centers of S-1 and C-l and at the 10- and








20-meter stations in S-2. Collected cores were returned to

the laboratory and stored at -20C. Preparation for analy-

sis consisted of thawing and homogenizing each section and

weighing 10 g wet sediment from each section into a 250-ml

Erlenmeyer flask for metals extraction. The remainder of

each section was placed in a tared crucible for determina-

tion of percent water, percent volatile solids, and sand

fractionation.

Percent water was determined by drying 24 hours at

1100C; percent volatile solids was determined by ashing the

dried, weighed sediment four hours at 450C in a muffle

furnace. After weighing, the ashed sediment samples were

wetted with a small amount of tap water and a few drops of

detergent and mixed thoroughly with a spatula. They were

then transferred to a U.S. Standard #200 mesh brass sieve

(nominal opening = 74pm) and wet-sieved with tap water until

all detergent had passed through the sieve. The sand frac-

tion remaining on the sieve was rinsed with distilled water

and washed into a tared aluminum weighing pan, dried at

1100C, and weighed to yield the percent sand in the dry

sediment. The combined percent silt and clay in each sec-

tion was determined by subtracting the sum of percent sand

plus percent volatile solids from 100 percent.

Metals extractable in 0.1 N HNO3 were extracted from

each core by adding 100 ml of 0.1 N HNO3 to the 10 g sedi-

ment in each flask, shaking 2 hours on a shaker table, al-

lowing to settle 2 hours, then vacuum-filtering through








acid-washed glass fiber filters (Gelman Type GF/A). The

filtrates were decanted into 125-ml polyethylene bottles

and refrigerated until analyzed by atomic absorption

spectroscopy.

The 0.1 N HNO3-extractable metals should include the

most mobile forms of the various metals in the sediments

(e.g., metals already in solution in the pore water, easily

dissolved salts such as chloride, carbonates, nitrates, and

some hydroxides, and metals available from ion-exchange

sites on clays and loosely sorbed on sand and other particu-

late matter). The 0.1 N HNO3-extracted sediments from core

S-1-0 (Dome S-l, 0 meters from center) were further ex-

tracted with aqua regia in an effort to distinguish be-

tween these relatively mobile metal species and the less

mobile forms such as metal sulfides, phosphates, and sili-

cates, metals incorporated into refractory organic materials

or clay structures, and metals occluded in amorphous oxides

and hydroxides.

The filters with the sediments collected on them were

transferred to 125-ml Kjeldahl flasks and digested with

50-ml aliquots of aqua regia until all organic were com-

pletely digested (4-6 hr). In most cases this required ad-

ditional aqua regia as the original aliquots distilled off

before complete digestion occurred. Digestions were con-

tinued until only 10-15 ml of liquid remained in each flask.

Digested samples were allowed to cool, diluted with dis-

tilled water, and again filtered through acid-washed glass








fiber filters to separate the sand, clay, and filter re-

mains from the liquid. The filtrates were diluted to 100 ml

preparatory to atomic absorption analysis, using flame atomi-

zation for Fe, Mn, Zn, Cu, Pb, and Ni, heated graphite tube

atomization for Cd, and the cold vapor technique for Hg.

4. Plant Tissue Analysis

Samples of duckweed (Lemma sp. and Spirodela sp.) and

cypress needles (Taxodium sp.) from the experimental domes

and several other locations were collected and analyzed dur-

ing April and May, 1976, to compare heavy metal levels in

macrophytes in the experimental domes with background levels.

Locations sampled included the three domes at the Owens-

Illinois site, the Austin Cary dome, the Santa Fe, Silver,

and Oklawaha Rivers, and a farm pond on the University of

Florida Swine Research Unit in Gainesville. Additional

samples from the Owens-Illinois site collected by investi-

gators involved with other aspects of the study were also

analyzed to supplement the data and to determine whether

metal concentrations in needles from individual cypress

trees have changed significantly since sewage application

began.

The latter group of samples had been prepared for analy-

sis by rinsing with tap water, drying at 600C, and grinding

in a ball mill. Samples collected during April and May,

1976, were prepared in the following manner: Both cypress

and duckweed samples were rinsed thoroughly with tap water.

Gross contamination (twigs, leaves, snails, spiders, etc.)








were removed from duckweed, and the fresh growths of cy-

press needles (needles occurring within the last six in-

ches of the branch) were stripped from each branch. Sam-

ples were then rinsed with deionized water and dried at

1000C for 24 hours. Two to four grams of dried sample (both

groups) were weighed into acid-washed, tared porcelain cru-

cibles and dry-ashed eight hours at 4500C. They were then

cooled in a dessicator, re-weighed to determine ash weight,

and the ash was taken up in a minimum volume of 4 M HNO3.

The crucible contents were transferred to 50-ml volumetric

flasks and diluted to volume with 0.1 M HNO3. The contents

of the flasks were mixed well and were left sitting over-

night to allow any insoluble residue to settle to the bot-

tom. The undisturbed supernatants were analyzed the follow-

ing day by AAS, using flame atomization for all elements

analyzed.

B. Laboratory Studies

Several laboratory investigations were designed to fur-

ther characterize the nature of organic compounds present

in sewage effluent and organic color compounds present in

cypress dome standing water, and to better elucidate the na-

ture, extent, and biological implications of heavy metal-

organic interactions in the environment.

1. Characterization of Natural and Sewage Organics

Water samples from four locations were collected for

concentration of dissolved organic and determination and

comparison of molecular weight distributions. The four








samples were from (1) sewage effluent from the Whitney mo-

bile home park, (2) the Austin Cary dome, (3) Lake Mize, a

highly colored oligotrophic lake in the Austin Cary Forest,

and (4) Newnan's Lake, a relatively uncolored eutrophic

lake east of Gainesville.

Two-liter to four-liter samples were obtained and re-

turned to the laboratory, where they were filtered through

glass fiber filters and then through 0.45pm membrane filters

to remove particulate matter. Filtered samples were freeze

concentrated to a total volume of 200 ml by the following

procedure: Three to four liters of sample were transferred

to a stainless steel beaker which was then set in an ethylene

glycol bath cooled to -200C. The sample was constantly

stirred at a slow rate until a clear, thick ice cake formed

on the beaker walls. The remaining liquid was transferred

to a smaller beaker and the operation was repeated until

the sample was concentrated to the desired volume. A 25-ml

aliquot of each sample was reserved for gel permeation

chromatography, and the remaining portion was reserved for

ultrafiltration. Both portions were frozen at -50C until

the respective procedures could be performed.

The gel permeation chromatographic procedure utilized

a 1.5 cm i.d. by 90 cm glass chromatography column (Pharmacia

Fine Chemicals Model K-15) and an automatic fraction collec-

tor (Gilson G.M.E.). The column was packed to the 70-cm

mark with Bio-Gel P-2, 100/200 mesh, or to the 85-cm mark

with Bio-Gel P-6, 100/200 mesh (Bio-Rad Laboratories,








Inc.). Column effluent was collected in 10-ml fractions at

flow rates of 0.87 ml/min (Bio-Gel P-2) or 0.5 ml/min (Bio-

Gel P-6). Several column volumes of distilled, deionized

water eluant were passed through each column to insure

proper packing of the gel bed before application of stand-

ards or samples. Both columns were calibrated with Blue

Dextran 2000 (MW z 2x106) and with methyl orange (MW = 327)

or sodium chloride to determine the excluded or void vol-

ume (V ) and the total column volume (Vt), respectively.

The internal volume of the gel (Vi) is the difference be-

tween Vt and Ve. Distilled, deionized water was used as

eluant. Absorbance of the collected fractions was measured

at 420 nm in 4-cm cells using a Beckman Model DB spectro-

photometer. Sodium was measured by AAS, and total organic

carbon (TOC) was measured using a Beckman Model 915 Total

Organic Carbon Analyzer equipped with a Model 865 Infrared

Analyzer.

Ultrafiltrations were performed using Amicon Diaflo

ultrafiltration membranes and a magnetically stirred Amicon

Ultrafiltration Cell, Model 52, under nitrogen pressures of

30-60 psi. The following three membrane types were used in

this work:

UM-2 (nominal MW cutoff = 1000)

UM-10 (nominal MW cutoff = 10,000)

XM-50 (nominal MW cutoff = 50,000)

Two series of ultrafiltrations were performed. In the

first series a sample of each organic concentrate was









ultrafiltered through a UM-2 membrane to partition the

sample into a low molecular weight fraction (MW < 1000)

and a high molecular weight fraction (MW > 1000) for use in

a duckweed metals uptake study. In the other series free

metal ions in the sewage and Austin Cary concentrates were

first removed by passing these samples through a 1.5 cm X

17 cm ion exchange column packed with a strong cation ex-

change resin (Dowex 50 X -8, 20/50 mesh) in the acid form.

The ion-exchanged samples were then sequentially ultra-

filtered through XM-50, UM-10, and UM-2 membranes to parti-

tion these two samples into four fractions with different

molecular weight ranges. In both series each fraction (ex-

cept the fractions with MW < 1000, i.e. the ultrafiltrates

passed by the UM-2 membranes) was concentrated by ultra-

filtration to a total volume of 30 ml. It was then diluted

1:1 with distilled, deionized water and re-concentrated to

a total volume of 40 ml in order to better de-salt the frac-

tion. This ultrafiltrate was added to the previous ultra-

filtrate from the sample before proceeding with the next

membrane in the sequence. All fractions were analyzed for

TOC, and the eight sewage and Austin Cary fractions from

the second series were analyzed for Cu and Cd by AAS.

These eight fractions were subsequently used in DPASV, ISE,

and ion exchange studies of organic interactions with copper

and cadmium.


______________________~








2. Heavy Metal Uptake by Duckweed

Two independent investigations were conducted to de-

termine the effects of sewage, EDTA, and natural color

organic on heavy metal uptake by duckweed. In the first

investigation, duckweed collected from dome C-l was cultured

two weeks in one-fourth strength Hutner's (1953) medium

under continuous fluorescent lighting of 6000 lux (560 ft.-

candles). Ten fronds of healthy duckweed were then trans-

ferred to each beaker in two series of 100-ml beakers con-

taining various modifications of Hutner's medium spiked

with 0 to 500 pg/1 Cu, and were incubated one week. Sur-

viving fronds were harvested, rinsed with distilled water,

and blotted dry. They were then weighed, ground in a tissue

grinder, extracted with 90% acetone, and analyzed for

chlorophyll-a by the trichromatic method (A.P.H.A. 1973).

The acetone was evaporated and the residue was digested

with 10 ml of 0.1 N HNO3 and 1 ml of 30% H202. The digested

samples were analyzed for Cu by AAS.

The second investigation utilized a growth medium

consisting of 0.45 pm membrane-filtered water from Lake

Magnolia, an oligotrophic lake about 25 mi northeast of

Gainesville, diluted 1:4 with tap water and spiked with

1.15 mg NH3-N//, 1.10 mg NO3-N//, and 1.21 mg PO4-P//.

The duckweed used in this investigation was collected from

the Santa Fe River, cleaned of snails, spiders, and debris,

and cultured for eight weeks in the Lake Magnolia medium

before setting up the experiments. Three times during that








period the healthy plantlets were transferred to fresh

medium and the old plantlets and medium were discarded.

Twenty fronds of duckweed were then transferred into

each beaker in four series of 100-ml beakers receiving

different combinations of Cu, Cd, Pb, or Zn with the low

and high molecular weight organic concentrates from the

first series of ultrafiltrations (see section 1). The

duckweed was incubated under a fluorescent light of 6,000

lux intensity on a 16 hr light, 8 hr dark cycle, and at the

end of one week the number of fronds in each beaker was

counted, noting the ratio of healthy/chlorotic fronds.

The fronds were then harvested, rinsed with distilled, de-

ionized water, and dried at 100C for 24 hr before weighing.

After weighing, the samples were dry-ashed 2 hr at 4500C,

and the ashes were transferred to 15- X 150-mm test tubes

with 10 ml of 0.1 M HNO3. Contents of the test tubes were

mixed well and allowed to settle overnight before analyzing

for metals by AAS.

3. Heavy Metal Release via Duckweed Decomposition

Duckweed was collected from dome S-l, cleaned in the

manner described in section A-4, and was killed by drying

48 hours at 500C. Five-gram portions were added to each of

two beakers containing 200 ml of glass fiber-filtered water

from dome C-l, adjusted to pH 6.8. Both beakers were cov-

ered with plastic film; one was maintained aerobic by

sparging with pre-moistened air, while the other was main-

tained anaerobic by sparging with pre-moistened nitrogen.








Ten-ml aqueous samples were taken prior to addition of

duckweed, 4 hr after addition, and at weekly intervals for

three weeks, and were frozen at -50C immediately after col-

lection. Samples were thawed at the end of the three-week

period and were filtered through acid-washed glass fiber

filters, and then through 0.45 pm membrane filters. Each

sample was transferred to a test tube and received 1 ml

concentrated, redistilled HNO3 and 1 ml 30% H202. They

were digested until the solutions cleared, diluted to 10 ml,

and analyzed for heavy metals by AAS.

4. Effects of Model Organic Compounds on DPASV Diagnostic
Parameters and ISE Response

A differential pulse anodic stripping voltammetry

(DPASV) study was conducted to examine the effects of four

organic compounds, chosen as models for natural organic,

on the diagnostic parameters i and E Instrumentation

consisted of a Princeton Applied Research Model 174 Polaro-

graphic Analyzer equipped with a hanging mercury drop elec-

trode (HMDE) and a Houston Instruments Omnigraphic X-Y Re-

corder used to record the output. Instrumental parameters

were adjusted for ASV analyses in the differential pulse

mode, with a voltage scan rate of +5 mv/sec, an applied

modulation amplitude of 50 mv, and a pulse rate of 1/sec.

Depositions were carried out at -0.4 v (Cu) or -0.9 v (Cd)

versus a saturated calomel electrode (SEC) for 60 sec (50

sec controlled stirring, 10 sec quiescent) before starting

the stripping step. The analytical procedure was as follows:








A 50-ml aliquot of distilled, deionized water was intro-

duced into the cell, and a 50-p. aliquot of 6 M acetate buf-

fer was added to provide a combination 6 mM acetate support-

ing electrolyte and pH 5.10 buffer. The solution was de-

aerated 30 min with vanadous chloride-scrubbed nitrogen to

remove all traces of oxygen, and a DPASV analysis was per-

formed to establish a baseline and to determine whether

the cell had been adequately cleaned between samples. A
_o
5 pZ aliquot of 102 M metal stock solution was added to

yield 10-6 M metal in the cell, and another DPASV scan was

performed. An aliquot of a 0.3 M model organic stock solu-

tion was added, the cell contents were mixed 5 min, then

the DPASV scan was repeated, and the pH was recorded. This

procedure was repeated with additional aliquots of ligand

stock until a range of ligand concentrations from about

1 yM to 3 mM had been investigated. The cell was then

emptied and cleaned by scrubbing with soapy water, acid

washing with 1:1 HNO3, and rinsing with distilled, de-

ionized water in preparation for the next sample. The pro-

cedure was continued until all four model compounds had

been investigated in combination both with Cu and with Cd.

An investigation examining the effects of the model

compounds on potentials developed by Cu and Cd ISE's uti-

lized the same experimental design as the DPASV investiga-

tion. The principal procedural differences were the elimi-

nation of the deaeration step and the use of an Orion Model

801 Digital pH Meter to monitor the potentials developed by


1








an Orion Model 94-29A copper ISE or an Orion Model 94-48A

cadmium ISE. All potentials were measured versus an Orion

Model 90-01 single junction reference electrode which has

the same electrochemical characteristics as a saturated

calomel electrode. After addition of the final aliquot of

ligand and measurement of the resulting potential, each

sample was preserved with 0.5 ml of HNO3 and was frozen

until analyzed for Cu or Cd by AAS.

5. Effects of Natural and Sewage Organics on ASV
Diagnostic Parameters and ISE Response

The eight molecular weight fractions partitioned from

the sewage and Austin Cary dome water concentrates (see

section 1 above) were utilized in DPASV and ISE investiga-

tions paralleling those with the model compounds. The

DPASV investigation differed from the model compound work

in the following respects: 1) Both Cu and Cd were added to

the solution in the analysis cell initially, and their inter-

actions with added organic were observed simultaneously by

using a plating potential of -1.0 v and scanning through

both the Cd and Cu DPASV peaks on each analysis. 2) Cu and
-7
Cd levels added initially were 107 M. 3) The ligands in-

vestigated were the eight organic fractions, and only two

concentrations, 10 mg TOC/Z and 20 mg TOC/e, were used.

4) After addition of the second ligand aliquot and the sub-

sequent DPASV analysis, two additional aliquots of Cu and
-7
Cd were added to raise the concentrations to 2 X 10 M

and 3 X 10-7 M, and DPASV scans were run at these levels.
and 3 X 10 M, and DPASV scans were run at these levels.








5) After the final DPASV scan the pH was measured and each

sample was acidified with 0.5 ml HNO3 and analyzed for total

soluble Cu and Cd by AAS.

The first ISE investigation with natural and sewage

organic was conducted in the same manner as the model

compound studies, except that only one ligand aliquot,

yielding a TOC value of 10 mg/, was added in each case.

A second ISE/AAS investigation addressed the question

of the liabilities of complexes formed between Cu or Cd and

the natural and sewage organic. Free Cu and Cd in equi-

librium with organic ligands in solutions identical to

those used in the first investigation were measured by ISE,

and total soluble Cu and Cd were measured by AAS without

acidifying the samples. Then 0.1 g Dowex 50 X -8 cation

exchange resin was added to each solution and mixed 15 min.

Free and total Cu and Cd were redetermined, and equilibrium

ratios of free metal to completed metal, Mf/Mc were cal-

culated and compared before and after addition of ion ex-

change resin.

6. Determination of Complexation Capacity of Model Organic
Compounds and Natural and Sewage Organics by Ion
Exchange Equilibrium

The final laboratory study utilized the ion exchange

equilibrium method developed by Crosser and Allen (1976) for

determination of complexation capacity of soluble ligands

(see Ch. II-B). The ligands investigated were the four

model organic compounds and the eight fractions from the

sewage and Austin Cary dome water concentrates. Several








preliminary runs were conducted to determine the proper

initial conditions to yield a range of equilibrium metal

concentrations amenable to AAS analysis. The procedures

and conditions resulting from the preliminary runs were as

follows: 0.2 g cation exchange resin in the hydrogen form

(Dowex 50 X -8) was added to each of ten 15 X 150 mm test

tubes, followed by 15 ml of a solution containing a 0.08 mM

acetate pH buffer and a 0.07 M NaNO3 ionic strength buffer.

A suitable aliquot of ligand was added to each tube to

yield a concentration of either 10 mg/ (natural and sewage

organics, 2 mM phenylalaninee, salicylic acid, ferulic

acid), or 0.2 mM (tannic acid), and the tubes were stop-

pered and shaken in a rotary test tube shaker for one hour.

Aliquots of Cu or Cd stock solutions were added to each

test tube to yield a range of initial metal concentrations

of 0.03 mM to 3.0 mM for runs with the model organic com-

pounds, or 0.003 mM to 0.3 mM for runs with the natural and

sewage organic. pH values were adjusted into the 5.5 to

6.5 range by adding microliter aliquots of 1.0 N NaOH and

1.0 N HNO3. The tubes were then re-stoppered and shaken

12 hrs. At the end of 12 hrs they were removed from the

shaker, pH values were determined, and total soluble Cu or

Cd concentrations were measured by AAS. The quantity of

metal taken up by the resin in each tube was calculated

from the difference between the initial and final concen-

trations, and isotherms were plotted for each metal-ligand

combination.















CHAPTER IV
RESULTS AND DISCUSSION

The presentation of results and discussion in this

chapter follows the same general outline and sequence used

in Ch. III, i.e., the two major subdivisions within the

chapter are field studies and laboratory studies, and

several investigations are described in each subdivision.

A. Field Studies

1. Routine Monitoring

Objectives of the routine monitoring efforts, in both

this and the earlier work by Klein (1976) were character-

ization of heavy metal concentrations in sewage applied to

the cypress domes and in standing waters within the domes,

and evaluation of possible heavy metal contamination of the

water table aquifer by seepage through the dome floors.

Table 6 lists the means, medians, and ranges of heavy

metal concentrations in the oxidation pond and treated

sewage effluents, and the standing waters in S-l, S-2, and

C-l, and Table 7 lists these statistics for the groundwater

pumped to C-l, and selected shallow wells over the period

from October, 1975 to August, 1976. This period included

seven sampling dates, and one to three locations in the

experimental domes were averaged together; therefore the

number of samples represented by each value in the tables

ranges from 7 to 21. With such small sample populations,

49











Table 6.


Aqueous heavy metal concentrations (ug/) at several sampling stations
at the Owens-Illinois Site--October, 1975 to August, 1976.


Package Plant Effluent
n Median R


anqe


Mean


Oxidation Pond
Median


MenMedian Ran


Effluent
Rangp


0.01-0.43
0-23
15-311
0-79
0.2-6.8
0-690
1.0-8.4
0-21


0.03
4
20
4.4
0.96
4.6
4.8
9.2


0.03
2
20
6.0
0.7
1.0
3.4
9.2


0.02-0.05
1-11
4-40
0-13
0.3-2.0
0-20
0.7-10.2
9.2-9.2


Sewage Dome 1** Sewage Dome 2**
Mean Median Range Mean Median Range
Fe* 0.47 0.38 0.11-1.10 0.30 0.24 0.08-0.64
Mn 18 18 5-40 12 11 2-31
Zn 16 17 3-31 20 17 5-60
Cu 9.1 4.1 0-61 7.6 6.0 0-27
Cd 1.7 0.7 0-10 2.6 0.7 0-28
Pb 7.6 1.5 0-39 8.6 1.9 0-39
Hg 14 10 1-53 16 6.0 0-88
Ni 8.0 6 2.8-17 11 5.7 2-36


Mea


Fe*
Mn
Zn
Cu
Cd
Pb
Hg
Ni


0.17
12
73
13
1.8
44
4.3
9.5


0.17
12
52
7.0
1.0
2.0
2.5
9











Table 6. Continued.


Groundwater Dome***
Mean Median


(C-l)
Range


Range Me an


Austin Cary Dome
Median


0.14
9
12
3.4
0.6
2.0
4.6
11


0.03-0.57
2-38
3-80
0-10
0.1-4.0
0-39
0.9-26
8-17


0.32
32
70
14
1.8
9.0
12
13


0.31
20
51
9.5
1.1
3.8
10
18


0.15-0.49
18-96
31-191
1.6-41
0.1-6.0
0.9-31
1.4-31
2.6-19


Values in mg/Z
** Averages of 3 stations
*** Averages of 2 stations


Fe*
Mn
Zn
Cu
Cd
Pb
Hg
Ni


0.18
11
21
3.6
1.1
7.6
6.2
12


Range


Mean











Table 7.


Aqueous heavy metal concentrations (pg/) in selected wells
Owens-Illinois Site--October, 1975 to August, 1976.


Well 30 (west of
Mean Median


C-1)
Range


Range Me an


Well 28 (east of
Median


0.35-9.89
55-22
12-222
0.0-13
0.2-3.0
0.1-24
0-34
5.5-18


1.19
8
346
12
3.4
15
7.7
18


1.19
8
36
14
0.8
14
3.8
11


0.44-1.93
0-15
16-1295
0.5-19
0.3-11.9
0.1-30
0-23
3-40


Groundwater Pumped to C-1 Well 8A (center of S-1)
Mean Median Range Mean Median Range
Fe* 0.11 0.05 0.02-0.44 3.55 3.48 2.95-4.29
Mn 4 3 1-12 10 10 7-14
Zn 43 13 1-40 16 12 2-40
Cu 9.1 9.0 2-20 6.1 6.0 0.2-16
Cd 5.0 1.0 0.1-22 0.6 0.2 0-3.3
Pb 12 3.6 0.4-42 8.0 2.3 0-29
Hg 3.6 4.0 0-6.2 20 11.1 1.2-80
Ni 17 14 12-24 6 4 4-10


at the


Fe*
Mn
Zn
Cu
Cd
Pb
Hg
Ni


4.57
18
65
5.2
1.5
8.7
9.8
10


4.20
19
22
1.0
1.8
5.0
5.5
8


C-1)


Range


Mean











Table 7. Continued.


Well 19 (west of S-1)
Median Range


Well 22 (east of S-1)
Mean Median


0.16-5.12
11-46
10-30
0-18
0-4.0
0-30
0.5-22
4-10


0.66
9
26
6.0
1.7
12
12
4


0.47
8
26
4.1
0.8
9.0
7.7
5


0.15-1.40
5-15
11-44
1.0-12
0-5.9
0-29
1.1-42
0-7


Mean


Fe*
Mn
Zn
Cu


1.81
26
18
5.0
1.0
10
6.3
7


0.68
22
16
2.9


0.6
2.7
2.5
8


Range










Table 7. Continued.


Well B-3 (north of S-2)
Mean Median Range


Mean


0.54-1.00
2-6
12-350
1.8-249
0.2-2
1.5-76
0.7-60
5-13


2.97
6
21
20
0.7
16
12
5


Well B-7 (south of S-2)
Median Range


2.34
5
15
12
0.9
7.2
8.8
6


1.17-6.51
3-8
8-48
0.2-75
0.2-1.0
0-47
0.2-28.8
0-10


Well B-9 (east of S-2) Well B-5 (west of S-2)
Mean Median Range Mean Median Range
Fe* 2.53 2.47 1.84-3.12 2.10 2.07 1.82-2.35
Mn 7 8 4-8 8 8 5-10
Zn 51 53 9-104 26 18 11-58
Cu 15 7 0.4-51 24 26 1-81
Cd 1.3 1.0 0.4-2.6 0.9 0.9 0.2-1.7
Pb 12 8.6 0-34 13 10 0-29
Hg 12 7.7 5.4-28 31 8.2 1.3-180
Ni 9 11 3-14 7 6 4-12

* Values in mg/


Fe*
Mn
Zn
Cu
Cd
Pb
Hg
Ni


0.73
3
90
77
1.0
32
26
9


0.67
3


58
33
1.0
16
27
7


.a








the median values present a more accurate description of

normal concentrations than do the means. Comparing these

with values listed in Table 1 (Ch. II) shows that, except

for Fe and Hg, the heavy metal concentrations in these

waters are within the range of typical concentrations for

fresh waters.

Most of the waters listed in Tables 6 and 7 are anoxic

or nearly anoxic, which accounts for the Fe concentrations

exceeding typical values in (aerobic) fresh waters; how-

ever the high Hg values are less easily explained. It

seems unlikely that the actual values are as high as those

reported here. Three potential sources of error in the

analytical procedure are: 1) contamination from the sample

bottles, 2) interference from non-atomic absorption, and

3) sorption of Hg on particulate matter in the samples.

Eight randomly selected sample bottles were leached for one

week with a 5 percent H2S04, 0.01 percent KMnO4 solution to

determine if bottle contamination was the source of the high

Hg values. Analyses of the leachates yielded apparent Hg

concentrations ranging from 0.24 pg/Z to 1.90 pg/t and

averaging 0.96 pg/t, thus bottle contamination is a probable

contributor to the anomalous results, but the amount of

contamination appears to be insufficient to fully explain

the error. Non-atomic absorption at the 253.7 nm analytical

wavelength from ketones and other volatile organic formed

during sample oxidation by permanganate (Parker 1972) may

also contribute to some of the high Hg values, and some








error may be due to Hg sorbed on particulate matter which

passed through the glass fiber filters used in sample

preparation (Logsdon and Symons 1973), but the relative

contributions of these two sources remain undetermined.

Although there appear to be some discernible differ-

ences in heavy metal concentrations between sites (e.g.,

oxidation pond effluent < sewage effluent; shallow ground-

waters < standing waters), the differences are small, and

since in most cases the values for Mn, Zn, Cu, Cd, Pb,

and Ni are at or near detection limits, it seems more

appropriate to observe that heavy metal concentrations at

all stations are quite low than to overemphasize any small

differences which may exist. This also holds true for the

three wells (#22, #8A, and #19) which transect S-1 in the

direction of groundwater flow and the wells (B-3, B-5, B-7,

and B-9) around the periphery of S-2. Although well 19,

downstream of S-1, has shown evidence of contamination due

to seepage from the sewage dome (high concentrations of

major cations, chlorides, and coliform bacteria), it has

not shown signs of heavy metal contamination during the

October, 1975 to August, 1976 period. This, in conjunction

with the data for the wells around S-2, indicates that no

significant heavy metal contamination of the shallow

groundwater aquifer occurred during that period.

Elucidation of any seasonal trends in the routine

monitoring data has been complicated by the very low heavy

metal concentrations present at all stations and by









variations during the three-year course of the larger

interdisciplinary investigation in sewage and groundwater

application rates to domes S-1 and C-1. Fairly high heavy

metal concentrations were observed at several stations dur-

ing the first few months of the project, but these were

attributed to mineralization effects of a forest fire which

swept the area in December, 1973 (Klein 1976). These high

initial concentrations decreased to values comparable to

those reported in Tables 6 and 7 within 12 months after the

fire, and have remained low for the duration of the project.

2. Sediment Characterization and Metal Profiles

Investigation of physical characteristics and metal

profiles in sediments in domes S-1, S-2, and C-l had the

following objectives: 1) to assess the downward mobility of

heavy metals in these sediments, 2) to determine whether

heavy metals are retained within the domes by sediments, or

whether they percolate through the floor of the domes into

the groundwater, and 3) to evaluate whether any significant

differences exist between the sewage domes and the ground-

water dome with respect to general stratigraphy or metal

concentrations in sediments.

Profiles of water, volatile solids, sand, and clay and

silt in the sediments of domes S-1, S-2, and C-1 presented

in Figures 4-7 show several common trends as well as some

characteristics peculiar to each dome. In all three domes

the water content and the volatile solids content of the

sediments is highest in the upper 6 cm and gradually





















EI-0
0 50% 100%


S9

S J -1E


E2-0


100%


CI-0
50% 100%


E2-12
0 50% 100%













CI-IO
0 50% 100%


Figure 4. Percent water profiles in sediment cores
from the Owens-Illinois site.


100%














100%


100%


a' 18
o

27'






(
0.
9-
9.

E 18

. 27


E2-20


CI-20






















EI-0
50% 100%













E2-0
50% 100%


EI-10
0 50% 100%
0 "

9

18-

27

36



E2-12
o 50% 100%


0I

18_


Cl-10


EI-20
50% 100%


E2-20
50% 100%


CI-20


Figure 5. Percent volatile solids profiles in
sediment cores from the Owens-Illinois
site.


. 9.
-
18-
E











I 9


r
ae


















EI-0
0 50% 100%
0

O l' "

S18I


S18
27 9

27


El-10
.50% 100%


E2-12
0 50% 100%


91

18


Cl-0
50% 100%


CI-10
0 50% 100%


Figure 6. Percent sand profiles in sediment cores
from the Owens-Illinois site.


El-20
50%


S18

.27
0
























27

36

45

E2-20
0 10% 50%


9.

18.


Figure 7.


Percent clay and silt profiles in sediment
cores from the Owens-Illinois site.


EI-IO


E2-0


50%


E2-10


50%








decreases with depth, stabilizing at 20 5 percent water

and less than 5 percent volatile solids at depths of 25-

50 cm. Dome S-1 exhibits lower values for percent water at

the 0-meter station and lower values for volatile solids at

all three stations than might be expected in surficial

sediments in systems receiving sewage. S-2 yielded the

highest values observed for these two parameters at the 0-

and 10-meter stations. The groundwater dome sediments con-

tain a layer at about 20-30 cm depth which has higher

water and volatile solids content. There is no evidence

for a similar stratum in the top 50 cm of either of the

sewage domes.

The sand profiles indicate that surficial sands present

in the dry areas of the Owens-Illinois site extend into and

through the upper 50 cm strata of the cypress domes. The

sand content of the sediments approaches 90 percent in the

lower portions of cores from all three domes. The vari-

ability of the sand content in the upper portions of the

cores is more pronounced, ranging from 50-75 percent in the

top 9 cm of S-1 cores, from 3-45 percent in the top 6 cm of

S-2 cores, and from 35-55 percent in the top 9 cm of C-l

cores. One exception was a very low values (9 percent) in

the first segment of core # C-1-10. Decreased sand content

at the 20-30 cm depth in the C-l cores also is evidence of

an organic-enriched stratum beneath that dome. The clay

and silt profiles show less pronounced differences between

the three domes. Interpretation of any subtle trends in








these profiles must be tempered by the realization that

errors in the volatile solids and sand determinations were

compounded in determining clay and silt by difference; how-

ever the profiles do show that the clay and silt fraction

of the sediments is generally about 20 percent throughout

S-1 and S-2 and slightly higher (-30 percent) in C-l.

The metal concentration/depth profiles for sediments

of the three domes are depicted in Figures 8-15, and Table 8

characterizes the concentrations presented in these figures.



Table 8. Representative values of low, average, and high
concentrations of 0.1 N HNO3-extractable metals
in sediments from the Owens-Illinois cypress
domes. (Concentrations in mg/kg dry weight)


Metal Low Values Average Values High Values

Cu <0.025 0.050-0.150 >0.250
Hg <0.050 0.050-0.200 >0.250
Ni <0.200 0.400-0.800 >1.00
Pb <0.100 0.100-0.500 >0.750
Fe <50 100-400 >500
Mn <0.250 0.250-1.50 >2.50
Zn <0.500 0.500-2.50 >5.00
Cd <0.010 0.010-0.040 >0.080



The values reported here are considerably lower than

those reported by Klein (1976) for surface sediments in the

same domes and for one core in S-1. The differences un-

doubtedly are due to the different methods used in preparing

samples for analysis. Klein dry-ashed the sediments at

3500C before dissolving in HNO3 and analyzing for metals--a

procedure which yields total metal concentrations present








without discriminating between forms, and which probably

includes some forms not extracted or inefficiently extracted

by the procedure used in this work. The values are also two

to four orders of magnitude lower than values reported by

Lund et al. (1976) for metals extracted from cores taken

from sewage disposal ponds, using 4M HNO3 extractant, and

are one-half to one-tenth as large as values reported by

Williams et al. (1975) for metals extracted from sediment

cores from seven lakes in upstate New York, using a mixed

HC1/H2SO4 extractant.

A duplicate core taken at the center of dome S-2

yielded very different water, volatile solids, and sand

profiles, and extracted metal concentrations were much

higher than the values listed in Table 8. Although a re-

view of the procedures, data, and calculations failed to

reveal a systematic error, because of the magnitude of the

differences between this core and the others, inclusion of

these values in the compilation of Table 8 would have dis-

torted the description of metal concentrations extracted

from the other cores. However, the metal profiles for this

core are presented in Figures 8-15. Two possible explana-

tions may account for the anomalous results for this core:

(1) The magnitude of the metal concentrations suggests that

a stronger acid concentration may have been inadvertently

used in the extractions, and (2) the anomalous water,

volatile solids, and sand profiles suggest that this may

be a non-representative core--a distinct possibility








considering the number of sediment-disturbing construction

activities (construction of a tower, a boardwalk, and a

one-room house) which have been concentrated at the center

of S-2.

Values obtained from the aqua regia extracts of core

S-1-0 also were excluded in developing Table 8. As de-

scribed earlier (Ch. III-A-3), the aqua regia extraction

following extraction with 0.1 N HNO3 was designed to dis-

tinguish between relatively immobile and mobile forms of

metals in the sediments. Inclusion of the aqua regia

extract values in Table 8 would have been inappropriate and

would have biased the results. However, it is appropriate

to compare the abundances of the two groups of metal forms,

and since only one core was extracted with both procedures

the most valid comparison is between the two sets of metal

profiles for core S-1-0. These profiles appear in the upper

left areas of Figures 8-15. Since the relationships between

the two groups of metal forms probably are similar in domes

S-1 and S-2, and perhaps to a lesser extent in C-l, the

following discussion might reasonably be applied to the

cores from those domes.

Given the anaerobic conditions and the high sand con-

tent of the S-1 sediments, it seems likely that enrichment

of the less mobile metal species in surficial sediments is

evidence of their prior incorporation into refractory

organic compounds or in situ deposition of insoluble salts.

Enriched concentrations of these forms at greater depths









may be evidence of previously deposited strata. From this

viewpoint the data in Table 9 indicate that Cu and Hg are

being deposited in fairly immobile forms in the surficial

sediments in S-1. Cadmium appears to be retained in the

surface sediments to a lesser extent. Mn is about equally

distributed between less mobile and more mobile forms at all

depths, indicating involvement in a dynamic equilibrium. Pb

and Ni are slightly enriched in less mobile forms at the

surface but show even greater enrichment in the deeper sedi-

ments. The less mobile Zn species are depleted with respect

to the more mobile forms in both the surface and deeper

sediments, while the less mobile Fe species are depleted at

the surface and enriched in deeper sediments.



Table 9. Distribution of metals between less mobile and
more mobile species in core S-1-0.

Metal Approximate ratio of less mobile/more mobile species
Surface Sediments Deeper Sediments

Cu 10 4
Hg 5 3
Cd 1 <1
Mn 1 1
Pb 2 4
Ni 1 3
Zn 0.2 0.5
Fe 0.2 5



The highest concentrations of 0.1 N HNO3-extractable

metal observed in this study are for iron (Figure 8). Con-

centrations in surface sediments frequently exceed 250 mg/kg,

decreasing to values less than 10 mg/kg at depths greater









SI-0,
0


region
200


12780


SI-0
0 100 200

0-
E 376
r9
C-


SI-10


0o

9
E
- 18
0 -


S2-0,# I
0 500 1000


9

18

27


S2-10
0 100 200 300
S77
9910


18


SI-20 S2-20
100 200 0 100 2i


CI-20


54-
Figure 8. Iron concentrations (mg/kg) in sediment
cores from the Owens-Illinois site.


Cl-0








than about 15-25 cm in S-1 and S-2, and 35-40 cm in C-l.

The 0.1 N HNO3-extractable Fe values in C-1 are somewhat

higher than in S-1 or S-2, probably reflecting local varia-

tions in geology. A band of higher concentrations at 18-

30 cm depth in C-1 interrupts the trend toward decreasing

concentrations with increasing depth and suggests that Fe

and other heavy metals may be strongly associated with the

high organic matter content of this stratum. Iron con-

centrations increased toward the centers of all three domes.

Manganese concentrations (Figure 9) are much lower

than Fe concentrations in these sediments, as is the case

for most sandy Florida soils. Manganese concentrations in

S-2 are generally higher than in S-1 or C-l, with peak

concentrations of 10-20 mg/kg at the surface. The lowest

values were observed in S-1, where surface sediments contain

0.5-2.5 mg Mn/kg. Concentrations in C-1 are intermediate

between these two, but persist to greater depths, with

peaks at the surface and at 20-30 cm depth.

Examination of Figure 10 reveals that 0.1 N HNO3-

extractable copper concentrations in all three domes are

about 0.1-0.2 mg/kg in the surface sediments and, with a

few exceptions, generally decrease with depth. Comparing

the 0, 10, and 20 m stations within each dome, it appears

that gradients toward higher copper concentrations at the

centers of the domes exist in S-1 and S-2, while the

reverse is true for C-1.








SI-0, Aqua regia
0 ID 2.0

S9

18



SI-0
0 1.0 2D


0

4 8-
Q


Sl-IO
0 1.0 2D
0- '

9"

18-


36.







0.

9
E
S18

27

36


S2-0,#2
0 5 10
S28.9
14.2
O~y",:


S2-10
ID 2.0


S2-20


SI-20


CI-0
1. 2.0



I


--- --


C1-10
1.0 2.0


Cl-20


Figure 9. Manganese concentrations (mg/kg) in sediment
cores from the Owens-Illinois site.









SI-0,Aquo region
0 0.5 1.0


SI-0
0 0.10 0.20


9.

I8


SI-10


E 9
J
. 18"
A)
a


S2-0,* 2
0 0.10 0.20
0 - -- -
0 1.33








9 19I 20
0

9

I8

27 7/-t32.2
2V


S2-10
0 0.10 C


SI-20
0 0.10 0.20
0 -

9-

E
S18"

=27
o
36

45


Cl-0


CI-IO
0.10 020


CI-20


Figure 10. Copper concentrations (mg/kg) in sediment
cores from the Owens-Illinois site.








The mercury profiles (Figure 11) show some variation

between the domes, with S-1 consistently having the lowest

values (generally <0.1 mg/kg), S-2 having slightly higher

values (median -0.125 mg/kg), with wide fluctuations with

depth. Concentrations in S-1 and S-2 decreased slightly

with increasing depth and with increasing distance from the

center. No spatial trends were discernable in C-l.

Nickel concentrations showed a pronounced gradient with

depth (Figure 12). Concentrations in the surficial sedi-

ments generally approach 1 mg/kg and rapidly decrease in

lower strata. Values exceeded 1 mg/kg in the upper 9 cm of

cores S-2-0, S-2-10, and C-1-20, and in the upper 3 cm of

Core C-1-10. Levels in C-l were slightly higher than in

S-1 and S-2, with a band of higher concentrations at depths

of 21-30 cm interrupting the concentration/depth gradient.

The lead profiles in Figure 13 are quite similar to

the nickel profiles. Concentrations in surface sediments

are about 1 mg/kg, as for Ni, but more values exceed 1 mg/kg

than for Ni. Also, there is evidence of a band of higher

concentrations in the 20-30 cm depths in C-1 which is more

pronounced at the 10 m and 20 m stations.

The cadmium and zinc profiles (Figures 14 and 15) are

similar except for the magnitudes of the concentrations.

Cadmium concentrations are the lowest of all the metals

measured, decreasing from about 0.1 mg/kg at the surface to

less than 0.03 mg/kg at depths below 9 cm. Zinc concentra-

tions are generally less than 1 mg/kg in the deeper









SI-0,Aqua region
0 0.5 1.0


9-

18-


S2-0, I
0 0.10 0.20
0 = 1.28

9 -73.94

18. ICB-8.19

27-







S2-10
0 0.10 0.20


CI-0
0 0.10 0.20




18
/--1 5.92
27

36-

45-


C1-10
0 0.10 0.20
0. /10.508
9
0//-.361
18

27I

36 I
/-Q0.374
45

54-


S2-20
0 0.10 0.20


Figure 11.


Mercury concentrations (mg/kg) in sediment
cores from the Owens-Illinois site.


SI-0


0
0


SI-20
0.10 0.20


CI-20
0.10 0.20











935
I





1//-o.3










SI-0, Aquo region
0 0,5 1.0




18-



SI-0
0 0.5 1.0
0

9-

18-


SI-IO
05 1.0
Sl-10

I


S2-10
0 0.5 1.0
I 1.45
9 -

18


S 1-20
0.5


Figure 12.


S2-20
0 0.5 1.0
S '
9-

18-


CI-20
0.5 1.0

2.08




3-


54-
Nickel concentrations (mg/kg) in sediment
cores from the Owens-Illinois site.


16.33


E
~ 182

3 27


36'


2.78










SI-0, Aquo regia
0 0.5 ID
0


9.
18 ff



SI-0
0 0.5 1.0


9. 1

18









SI-10


SI-20
0 0.5 ID
0 1 2.12

9-

18-

27-

36


S2-0,r 2
0.5


S2-0,1 I
0 5
0I

9

18

2-7 z 7.30
9.92







S2-10
0 0.5 1.0
0 12.1
I/ 3.76
9. .41

18


S2-20


Cl-0


C1-10


Cl-20
0.5


54J

Figure 13. Lead concentrations (mg/kg) in sediment
cores from the Owens-Illinois site.










SI-O, Aqua region


10.847


SI-0


S9.
o.
o 18


Sl-10


Sl-20
0 0.05 0.10


9

E 18.
0 I


27
a-
36


45-


Figure 14.


S2-0, I
0 0.10 0.20
0 /Jo10.431

9

18- I

27-







S2-10
0 0.05 0.10
0 //-13.86

9

18


S2-20
0.05


Cl-0
0 0.05 0.10


9

18

27

36

45


0.10


CI-20


Cadmium concentrations (mg/kg) in sediment
cores from the Owens-Illinois site.










SI-0, Aquo region


p 5


0
0-




I,






S9
0



-
- 18 -


SI-10


SI-20


S2-20


Figure 15. Zinc concentrations in sediment cores
from the Owens-Illinois site.


Cl-0









sediments, increasing to values of 2-25 mg/kg at the surface,

with a few samples exceeding this range. There is a slight

trend toward decreasing concentrations toward the edges of

all three domes.

In summary, the volatile solids (organic) content and

the sand content of the sediments in S-1 and S-2 are in-

versely related, with organic content decreasing the depth,

while sand content generally increases from 10 to 30 percent

at the surface to 80 to 90 percent (dry weight basis) at

depths 25 cm. Dome C-1 has an organic-enriched stratum

extending from 18 to 30 cm depth. Concentrations of 0.1

M HNO3-extractable heavy metals in these cores are much

lower than values reported in the literature for heavy

metals extracted from sediments of sewage disposal ponds

with stronger acids; the differences probably are attribut-

able to the more rigorous extractions utilized in those

studies, differences in heavy metal content of the sewage

applied and the lengths of application, and the high sand

content of the soils at the Owens-Illinois site. Metal

concentrations are generally higher in the surface sediments

and decrease rapidly with depth. Results of aqua regia

extractions following the 0.1 M HNO3 extractions indicate

that Cu and Hg are immobilized in the surface sediments,

while Fe, Ni, and Pb are immobilized in somewhat deepter

sediments, and Mn, Zn, and Cd are approximately equally

distributed between less mobile and more mobile forms at all

depths.








3. Plant Tissue Analyses

Duckweed (Lemma sp. and Spirodela sp.) is an important

component of the macrophyte community in the sewage domes.

It appeared shortly after sewage was applied and grew

rapidly, forming a thick floating mat covering most of the

standing water surface. It also appeared in lesser abun-

dance in C-l, and in early summer, 1975, was introduced into

the Austin Cary dome, possibly being transported there on

equipment and waders used at both the Owens-Illinois and

Austin Cary sites. Duckweed is a relatively short-lived

macrophyte, having an average life span of about three

weeks. Combined with its presence throughout the year,

this fact makes duckweed a potentially significant factor

in heavy metal cycling in the domes. Similarly, the domi-

nance of cypress (Taxodium sp.) among macrophytes and its

deciduous nature make cypress another potentially important

component in heavy metal cycling within the domes.

Table 10 lists heavy metal concentrations in duckweed

collected from several locations during April and May, 1976.

Two principal conclusions can be drawn from these data:

1) heavy metal levels in duckweed from the sewage domes are

not significantly different from concentrations in duckweed

from other locations, and 2) duckweed accomplishes an ap-

proximately 1000-fold biomagnification of heavy metal levels

in plant tissue as compared to solution concentrations

(Table 6). The tissue concentrations and biomagnification

factors observed in this work are comparable to values


















Table 10. Heavy metal concentrations (mg/kg dry wt.) in duckweed--April and May,
1976.


Location

Sewage dome 1
Sewage dome 2
Groundwater dome
Santa Fe River
U.F. Swine Unit Pond
Austin Cary Dome


Number of
Samples

7
4
4
2
1
1


Fe

463
625
359
387
818
620


Mn

220
262
378
846
195
506


Zn

205
103
71.7
104
40.6
1670


Cu

11.6
8.33
3.21
66.1
4.49
54.9


Cd

3.37
1.15
0.92
2.63
0.63
3.32


Pb

10.0
6.89
6.76
6.58
47.0
6.14


Ni

4.21
5.62
2.63
6.58
2.28
5.36








reported by Hutchinson and Czyrska (1975) for duckweed

from 23 Canadian ponds.

Heavy metal concentrations in cypress needles collected

during May and June, 1976, are presented in Table 11. Once

again, there are no significant differences in concentra-

tions in samples from the sewage domes and other locations.

Comparison of these values with the values in Table 10

shows that heavy metal levels in cypress needles are about

1 to 1.5 orders of magnitude less than in duckweed. Closer

examination of the data from which Table 11 was compiled

showed that the differences between trees within a dome

frequently were as large as the differences between domes,

and examination of data for heavy metal levels in needles

collected from the same trees in October, 1974, June, 1975,

November, 1975, and June, 1976 revealed no temporal trends,

even when the comparisons were between samples collected

during the same growing seasons in the two years (i.e.,

comparison of the data from June, 1975 and June, 1976, and

comparison of October, 1974 and November, 1975).

B. Laboratory Studies

1. Characterization of Natural and Sewage Organics

In view of the problems of apparent sorption of metal

ions by Sephadex polydextran gels reported by Klein (1976)

and Allen (1975), a polyacrylamide gel (Bio-Gel ) advertised

as being "biologically inert" and as having the "lowest

level of charged sites, 2 p-eq/dry g," (Bio-Rad Laboratories

1975) of any commercially available gels was selected for


















Table 11. Heavy metal concentrations (mg/kg dry wt.) in cypress needles--May and
June, 1976.



Number of
Location Samples Fe Mn Zn Cu Cd Pb Ni

Sewage dome 1 4 24.4 14.5 13.0 1.59 0.141 7.31 0.866
Sewage dome 2 2 37.7 43.4 16.0 1.81 0.271 12.4 1.33
Groundwater dome 2 19.2 22.7 13.8 1.62 0.372 9.60 1.66
Silver River 1 21.6 2.96 11.4 2.14 0.204 2.28 1.34
Santa Fe River 2 14.3 45.2 27.3 0.49 0.185 1.44 0.898
Oklawaha River 2 26.8 26.7 13.0 1.35 0.294 1.78 1.03
Austin Cary Dome 4 15.7 17.9 18.8 1.37 0.391 10.5 1.45








use in characterizing the molecular weight distributions of

natural and sewage organic. Ultrafiltration was also uti-

lized for this purpose, using Amicon Diaflo filter types

UM-2, UM-10, and XM-50. Diaflo membranes are selectively

permeable membranes "cast from non-cellulosic polymer

solutions, which reject solutes larger than their (molecular

weight) cut-off, passing microsolutes freely" (Amicon 1975).

Type UM membranes "possess some free ionic sites, but ex-

hibit a net neutral charge", while type XM membranes are

non-ionic (Amicon 1975).

All four organic concentrates prepared as described in

Ch. III-B-1 were chromatographed with Bio-Gel P-2, which

separates molecules in the 100-1800 Dalton range, and the

sewage and Austin Cary concentrates were chromatographed

with Bio-Gel P-6, which functions in the 1000-6000 Dalton

range. The sewage and Austin Cary concentrates were also

fractionated by ultrafiltration.

Chromatograms obtained with Bio-Gel P-2 using distilled

water eluant (Figures 16 and 17) show that this gel appar-

ently interacts strongly with the organic color molecules.

The principal color peaks eluted after the low molecular

weight standard, NaCI, and the principal TOC peaks eluted

in the same volume as the low MW standard. Small color and

TOC peaks were observed at Ve, corresponding to compounds

with MW > 1800. The amount of interaction might be de-

creased by using a higher ionic strength eluant, but it





83




300 150
AUSTIN CARY DOME



_0 COLOR
200- / -100 -
TOC \
o I E
St E
S- I --
o I
0 O- I
) 100- I -50
< I


0' 0

0 l l r l la ir I li a 0
0 t 50 t 100 150 200
Ve Vt
ELUTION VOLUME, ml


.100- 50
NEWNANS LAKE


z N
< ~TOC-,I \ COLOR E
m .050- -25

i\ O
S I J-



..000 0
0 t 50 ( 100 150 200
Ve Vt
ELUTION VOLUME, ml


Figure 16. Color and TOC profiles for organic
concentrates from the Austin Cary
dome and Newnans Lake eluted through
Bio-Gel P-2.





84


300 310 30
LAKE MIZE i
I'
I -



S_ TOC-- l -COLOR

o= '
u -







0 50 100 150 200
V, Vt
ELUTION VOLUME, mi

.030 15
SEWAGE ,




.020- 10
I I I I I I I I I I I I I I I I I I I
O50 I 15 2o0
a .


0 0
M .010- 5






0 2 50 100 150 200
Ve Vt
ELUTION VOLUME, ml

Figure 17. Color and TOC profiles for organic
concentrates from Lake Mize and
sewage eluted through Bio-Gel P-2.
(Note scale change for low color
sewage concentrate.)
sewage concentrate. )









certainly obscures any useful information which might be

obtained using deionized water eluant.

The chromatogram of Austin Cary concentrate obtained

with Bio-Gel P-6 using distilled water eluant (Figure 18)

shows less evidence of interaction with the gel. Both the

color and TOC peaks eluted within the range bracketed by

Ve and Vt, with the dominant peaks occurring at molecular

weights of about 3000 to 4000 Daltons, Dilution of the

sewage concentrate in the Bio-Gel P-6 column obscured any

significant molecular weight fractionation.

Since the chromatograms in Figures 16, 17, and 18 show

that Bio-Gel P-2 interacts strongly with organic color

compounds and that Bio-Gel P-6 produces very little molecu-

lar weight fractionation of these compounds, further efforts

to use these gels to characterize natural and sewage organ-

ics and heavy metal-organic interactions seemed likely to

prove fruitless. Therefore, a greater effort was concen-

trated on using ultrafiltration techniques for these purposes.

The molecular weight distributions obtained by ultra-

filtration (Figure 19) indicate nearly opposite distributions

for the sewage and natural organic. About 66 percent of

the sewage organic is in the lowest molecular weight

fraction, while 50 percent of the natural organic is in

the highest molecular weight fraction, and another 35 per-

cent is in the second highest fraction.

Table 12 presents results of Cu and Cd analyses of

these fractions. Since the concentrates were passed










.100- 100
AUSTIN CARY


%._ TOC
.050- COLOR -50


\"J


0 50 100 150 200
Ve Vt
ELUTION- VOLUME, ml


ELUTION VOLUME, ml


Figure 18. Color and TOC profiles for organic
concentrates from the Austin Cary
dome and sewage eluted through Bio-
Gel P-6.




University of Florida Home Page
© 2004 - 2010 University of Florida George A. Smathers Libraries.
All rights reserved.

Acceptable Use, Copyright, and Disclaimer Statement
Last updated October 10, 2010 - - mvs