United States
Environmental Protection Health Effects Research
Aaencv Laboratory
Cincinnati OH 45268
EPA-600/1-84-030 Jan. 1985
Health Effects of
i
Land Treatment:
Toxicological
Cd
N03
CMC
Pb
CCI2 = CHCI
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EPA-600/1-84-030
January 1985
Health Effects of Land Treatment:
Toxicological
by
Norman Edward Kowal
Toxicology and Microbiology Division
Health Effects Research Laboratory
Cincinnati, OH 45268
HEALTH EFFECTS RESEARCH LABORATORY
OFFICE OF RESEARCH AND DEVELOPMENT
U.S. ENVIRONMENTAL PROTECTION AGENCY
RESEARCH TRIANGLE PARK, NC 27711
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DISCLAIMER
This report has been reviewed by the Health Effects Research Laboratory, U.S.
Environmental Protection Agency, and approved for publication. Mention of trade
names or commercial products does not constitute endorsement or recommendation
for use.
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FOREWORD
"I he many benefits of our modern, developing, industrial society are accompanied
by certain hazards. Careful assessment of the relative risk of existing and new-
man-made environmental hazards is necessary for the establishment of sound
regulatory policy. These regulations serve to enhance the quality of our environment
in order to promote the public health and welfare and the productive capacity of our
Nation's population.
The complexities ofenvironmental problems originate in the deep interdependent
relationships between the various physical and biological segments of man's natural
and social world. Solutions to these environmental problems require an integrated
program of research and development using input from a number ofdisciplines.The
Health Effects Research Laboratory, Research Triangle Park, NC and Cincinnati,
OH, conducts a coordinated environmental health research program in toxicology.
epidemiology, and clinical studies using human volunteer subjects. Wide ranges of
pollutants known or suspected to cause health problems are studied. The research
focuses on air pollutants, water pollutants, toxic substances, hazardous wastes,
pesticides, and nonionizing radiation. The laboratory participates in the development
and revision of air and water quality criteria and health assessment documents on
pollutants for which regulatory actions are being considered. Direct support to the
regulatory function of the Agency is provided in the form of expert testimony and
preparation of affidavits as well as expert advice to the Administrator to assure the
adequacy of environmental regulatory decisions involving the protection of the
health and welfare of all U.S. inhabitants.
This report provides a general appraisal of the impact of toxicological
contaminants in wastewater when applied to land. It is assumed that only a minimum
of preapplication treatment is given so that the land itself serves as part of the
treatment system. With a better understanding of such factors as the toxic substances
present in wastewater, decomposition rates, and toxicology, more informed
decisions may be made on proper management practices necessary to protect public
health in the community.
F Gordon Hueter, Ph.D.
Director
Health Effects Research Laboratory
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ABSTRACT
The potential health effects arising from the land treatment of wastewater are
examined, and an appraisal of these effects made. The agents, or pollutants, of
concern from a health effects viewpoint are divided into the categories of pathogens
and toxic substances. Only the latter are considered in this volume, the former having
been discussed in a previous volume. The toxic substances include organics, trace
elements, nitrates, and sodium. These agents form the basis of the main sections of
this report.
For each agent of concern the types and levels commonly found in municipal
wastewater and the efficiency of preapplication treatment (usually stabilization
pond) are briefly reviewed. A discussion of the levels, behavior, and survival of the
agent in the medium or route of potential human exposure, i.e., aerosols, surface soil
and plants, subsurface soil and groundwater, and animals, follows as appropriate.
Finally, conclusions and research needs are presented.
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CONTENTS
Page
Foreword iii
Abstract iv
Tables vi
Acknowledgment vii
1. Introduction 1
2. Organics 3
3. Trace Elements 19
4. Nitrates 28
5. Sodium 32
6. Comparison with Conventional Systems 33
7. General Conclusions 34
References 38
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TABLES
Page
I. Most Frequently Detected Priority Organics in Raw Municipal
Wastewater 5
2. Air and Wastewater Concentrations at Muskegon Land
Treatment System 8
3. Common Types of Chemical Transformations in the
Environment 9
4. Biodegradability of Priority Organic Compounds 12
5. Removal Efficiencies of Refractory Volatile Organics at 23rd Avenue
Project Rapid Infiltration Site 15
6. Removal of Volatile Organics at a Prototype Slow Rate Land
Treatment Site 16
7. Concentration of Trace Elements in Untreated Municipal Wastewater
in U.S. and Recommended Irrigation Water Criteria 21
8. Annual Input of Trace Elements and Years of Land Treatment
Required to Exceed Recommended Cumulative
Input Limits 22
9. Cadmium Uptake by Corn During and After Wastewater
Sludge Application 22
10. Wastewater, Soil, and Pasture Plant Levels of Toxic Trace
Elements at Werribee Farm 24
11. Toxic Trace Element Concentrations in Cattle Liver and Kidney at
Werribee Farm 25
12. Cadmium Concentration in Foods and Calculated Dietary
Intake 26
13. Wastewater and Groundwater Nitrogen at Slow-Rate Land
Treatment Sites 29
14. Wastewater and Groundwater Nitrogen at Rapid-Infiltration
Land Treatment Sites 31
15. Survival Times of Pathogens on Soil and Plants 31
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ACKNOWLEDGMENT
The editorial and scientific contributions of Herbert R. Pahren are gratefully
acknowledged.
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INTRODUCTION
It is the purpose of this report to examine the potential health effects of land
treatment, and to provide an appraisal of these effects.The agents, or pollutants, of
concern from a health effects viewpoint can be divided into the two broad categories
of pathogens and toxic substances. The pathogens have been covered in an earlier
report, "Health Effects of Land Treatment: Microbiological" (Kowal 1982). That
report also contains more general introductory material on land treatment, and is to
be considered a companion report to this one. The pathogens include bacteria (e.g.,
Salmonella and Shigel/a), viruses (i.e., enteroviruses, hepatitis virus, adenoviruses,
rotaviruses, and Norwalk-like agents), protozoa (e.g., Enianweba&nA Ciardia), and
helminths (or worms, e.g., Ascarii,, Trichurix, and Toxocara), The protozoa and
helminths are oft en grouped together under the term, "parasites," alt hough in reality
all the pathogens are parasites. The toxic substances include organics, trace elements
(or heavy metals, e.g., cadmium and lead), nitrates, and sodium. Nitrates and sodium
are usually not viewed as "toxic" substances, but are here so considered because of
their potential hematological and long-term cardiovascular effects when present in
water supplies at high levels. These agents form the basis of the main sections of this
report. The major health effects of these agents are listed below:
Agent (Pollutant)
Pathogens
Toxic
Substances
Bacteria
Viruses
Protozoa
Helminths
Organics
Trace Elements
Nitrates
Sodium
Health Effect
Infection, Disease
• Hypersensitivity
• Acute Toxicity
- Mutagenesis and
Carcinogenesis
- Teratogenesis
• Other Chronic Effects
(cardiovascular, immunological,
hematological, neurological, etc.)
For each agent of concern the types and levels commonly found in municipal
wastewater and the efficiency of preapplication treatment (usually stabilization
pond) are briefly reviewed. A discussion of the levels, behavior, and survival of the
agent in the medium or route of potential human exposure, i.e., aerosols, surfacesoil
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and plants, subsurface soil and groundwater, and animals, follows as appropriate.
Finally, conclusions and research needs are presented.
Surface water pollution from land treatment site runoff is not considered since
proper system design should prevent direct runoff to surface waters (Sorber and
Outer 1975, Reed 1979, USEPA 1981). Surface discharge of overland flow effluent
may have similar consequences to those of conventional treatment, but little is
known in this area since examples are so few.
The present volume is devoted to the toxic substances. For a discussion of the
pathogens see Kowal (1982).
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ORGANICS
The potential health effects of toxic organic compounds are myriad. Systems
affected range from the dermatological to the nervous to the subcellular, and effects
produced range from rash to motor dysfunction to cancer. The degree of toxicity of
organic compounds varies widely, from essentially harmless (e.g., most carbo-
hydrates) to moderately toxic (e.g., most alcohols) to extremely toxic (e.g.,
aflatoxins).
A glance at the current edition of The Merck Index will reveal that the number of
organic compounds described thus far is almost unfathomable. Nearly any of these
may appear in wastewater. depending upon its sources. Thus, the discussion below
must be perforce rather general, and the presence of any particular toxic organic in
high concentration in the wastewater may require a site-specific evaluation of
potential health effects.
Types and Levels in Wastewater
Most common organics in domestic wastewater derive from feces, urine, paper
products, food wastes, detergents, and skin excretions and contaminants (from
bathing). In medium-strength sewage (700 ppm solids content), organics make up
about 75% of the suspended solids and about 40% of the filterable solids (colloidal
and dissolved), consisting primarily of proteins (40-60%), carbohydrates (25-50%),
and fats and oils (10%) (Metcalf, and Eddy 1972). After secondary treatment, the
more refractory and high molecular-weight organics predominate, e.g., fulvic acid,
humic acid, and hymathomelanic acid (Chang and Page 1978). In general, however,
the chemical nature of domestic wastewater remains poorly characterized.
Although most of the organics found in domestic wastewater are probably
harmless in a land treatment context, it has recently been found that fecal material
commonly contains mutagens. Thus, there is evidence that one of the causes of
colorectal cancer is the presence of carcinogens or co-carcinogens produced by the
bacterial degradation in the gut of bile acids or cholesterol (Thornton 1981). The
mutagenicity of feces can be increased by anaerobic incubation and by the presence
of bile and bile acids (VanTasselle/a/. 1982), and the level of mutagenicity generally
is lower in vegetarians than non-vegetarians (Kuhnlein el al. 1981). High levels of
chromosome-breaking mutagenic activity have also been found in the feces of
animals—dog, otter, gull, cow, horse, sheep, chicken, and goose (Stich el al. 1980).
The chemical nature of the fecal mutagens is unknown. In the case of the latteranimal
mutagens, evidence suggests that at least part of the mutagenic action is due to
hydrogen peroxide and the ensuing radicals which can be formed during oxidation of
many organic compounds (Stich et al. 1980).
In the State of Illinois ten domestic and industrial secondary effluents were
examined for mutagenicity by Johnston el al. (1982), with the results that all ten
effluents assayed showed significant mutagenicty. Mutagenic activity per unit volume
of effluent varied over a 4,500-fold range, and toxicity varied over a 120-fold range.
Selective extraction of whole effluents appeared to unmask mutagenic activity,
probably by separating mutagens and substances that interfere with the mutagen
assay. In several effluents there was evidence of several mutagenic compounds
present, and it appeared that the mutagens were predominantly nonpolar, neutral
compounds. There was no obvious influence of disinfection by chlorination on the
effluent mutagenicity. in spite of the fact that one would expect many mutagens to be
3
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formed by the action of chlorine on humic substances and other organics found in
wastewater.
The major contributors of toxic organics to municipal wastewaters are usually
assumed to be industrial discharges. However, household wastewater discharge may
represent an important contributor since many consumer products in daily use
contain toxic substances. A recent study (Hathaway 1980) identified consumer
products containing toxic compounds on EPA's list of 129 priority pollutants, which
may eventually end up in wastewater. The most frequently used products are cleaning
agents and cosmetics, containing solvents and heavy metals as main ingredients.
Next are deodorizers and disinfectants, containing naphthalene, phenol, and
chlorophenols. Discarded into wastewater infrequently, but in large volumes, are
pesticides, laundry products, paint products, polishes, and preservatives. The
organic priority pollutants most frequently used and discharged into domestic
wastewater were predicted to be the following:
benzene naphthalene
phenol toluene
2,4,6-trichlorophenol diethylphthalate
2-chlorophenol dimethylphthalate
1,2-dichlorobenzene trichloroethylene
1,4-dichlorobenzene aldrin
1,1,1-trichloroethane dieldrin
Because of the difficulty of analysis of complex mixtures, it has only recently been
possible to measure the actual levels of organics in wastewater using advanced
methods of extraction, gas and other chromatography, mass spectrometry, and
computer analysis. The U. S. Environmental Protection Agency has sponsored two
extensive surveys of the types and levels of priority pollutants in municipal
wastewaters, which, of course, result from both domestic and industrial discharges.
Thefirst(DeWallee/a/. 1981), supported by the Municipal Environmental Research
Laboratory in Cincinnati, covered 25 cities located throughout the United States,
and the second (Feiler 1980), supported by the Effluent Guidelines Division in
Washington, D.C., covered 40 cities, the results from 20 of which are reported in the
cited reference.
In the 25-city survey (DeWalle et al. ! 981) most of the 24-hour composite samples
of raw wastewaters contained a total of less than 1 mg/1 of priority organics, and the
numbers of compounds detected clustered between 20 and 50. In the 40-city survey
(Burns and Roe 1982) six days of 24-hour sampling was completed, and the samples
from 20 cities were analyzed. The priority organics detected in at least 50% of the
samples analyzed in either survey are listed, together with their concentrations, in
Table 1.
Comparison of the results of the two surveys with the list of organic priority
pollutants most likely to be discharged into domestic wastewater, reveals consider-
able overlap, and gives one some confidence that these two studies have yielded a
reasonable characterization of the priority organics, in municipal wastewater, at least
of those identifiable by modern methods. The broad range of concentrations detected
among the samples, however, suggests that wastewater applied to land should be
regularly monitored for toxic organics. This measure is emphasized by the occasional
discharge of toxic substances into municipal wastewater systems with resulting
medical effects in treatment plant workers, such as the recent hexachlorocyclo-
pentadiene episode in Louisville, Kentucky (Kominsky et al. 1980).
Preapplication Treatment
Temporary storage ponds are usually used with land treatment systems because of
the need to (1) remove grit, organic solids, and grease so as to prevent fouling of
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Table 1. Most Frequently Detected Priority Organics in Raw Municipal Wastewater
Compound
Phenol
1 ,1,1-Trichloroethane
Trichloroethylene
Tetrachloroethylene
Ethylbenzene
Trichloromethane (Chloroform)
Diethylphthalate
Di-n-butylphthalate
Toluene
Dichloromethane
Bis(2-ethylhexyl)phthalate
Naphthalene
1 ,4-Dichlorobenzene
Phenanthrene
Benzene
Heptachlor
Butylbenzylphthalate
BHC-G (Lindane)
1 ,2-Dichlorobenzene
Dimethylphthalate
DeWalle et al.
Detection
Frequency (%)
94
94
94
94
94
94
91
91
90
90
89
86
83
83
79
77
77
71
69
66
1981
Concentration
Range (fjg/\)
0.90-2440.00
0.40- 97.50
0.90-1553.00
1.50- 385.10
0.20- 304.40
0.25- 73.10
1.34- 290.00
0.26- 123.00
0.70- 795.00
0.50- 666.10
0.06- 117.00
1.25- 291.00
1.70- 119.00
0.20- 49.50
0.26- 243.00
0.30- 37.00
1.10- 237.00
0.05- 11.20
0.78- 703.00
0.09- 114.00
Burns and Roe
Detection
Frequency (%)
79
85
90
95
80
91
53
64
96
92
92
49
17
20
61
5
57
26
23
11
1982
Concentration
Range (jjg/\)
1-1,400
1 -30,000
1-1,800
1-5,700
1-730
1-430
1-42
1-140
2-1,300
1 -49,000
2-670
1-150
2-200
1-93
1-1,560
0.08-0.50
2-560
0.02-3.9
1-440
1-110
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Table 1. (Continued)
DeWalle et al. 1981
Compound
BHC-D
Dieldrin
1 ,3-Dichlorobenzene
BHC-A
DDT
Di-n-octylphthalate
1,1-Dichloroethane
1,2-Dichloroethane
ODD
Anthracene
Aldrin
Endosulfan-B
1 ,2-Trans-dichloroethylene
Detection
Frequency (%)
63
63
60
60
60
57
55
55
54
51
51
51
20
Concentration
Range (/yg/l)
0.01-
0.02-
0.08-
0.01-
0.10-
0.31-
0.20-
5.10
4.40
548.00
2.90
24.00
15.50
3.60
0.20-3950.00
0.05-
0.04-
0.02-
0.20-
0.20-
10.00
36.80
2.00
8.80
45.30
Burns and
Detection
Frequency (%)
3
1
7
8
<1
7
31
15
1
18
1
—
62
Roe 1982
Concentration
Range Oug/l)
0.10-1.4
0.03-0.04
2-270
0.02-4.4
1.2
2-210
1-24
1 -76,000
0.31-0.77
1-93
0.03-5
—
1-200
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wastewater distribution system components, (2) provide storage during winter when
crop uptake Tor nitrogen control is minimal and the soil biota may be inaccessible
and or relatively inactive, (3) provide storage during periods of precipitation, and (4)
equali/e wastewater flows. Storage ponds also serve a treatment function, as was
illustrated in the earlier report, on microbiology.
One of the few land treatment sites where the fate of organics has been followed
from input to output is that at Muskegon, Michigan (Demirjiane/a/. 1983). Here the
treatment train for about 30 MOD of combined industrial (over 70%, including
several organic chemical manufacturers) and domestic wastewater consists of 8 acres
of aerated lagoons, 850 acres of storage lagoons, and 5,500 acres of sprayirrigated
farmland, which is drained by subsurface (5-12 feet) tiles. The detention time of the
raw wastewater in the aerated lagoons is approximately 36 hours, and that in the
storage lagoons is variable, depending upon the season of the year.
The influent to the Muskegon system generally contains 20-30 priority organic
pollutants and 15-25 additional organic compounds. The aerated lagoons remove
over 90% of most compounds by stripping of the volatiles, over 99% of toluene and
acetone for example. Data have shown variable removal rates for organics in the
storage lagoons, sometimes quite poor, but organics average below 10/ug/l for most
compounds before spray irrigation. Removals for most compounds are better during
summer months than winter.
Aerosols
While most of the nonvolatile organics in wastewater at a land treatment site
would be expected to enter the soil, much ofthe volatile organics would probably be
released to the air. The fraction entering the air would depend, of course, upon the
processes, detention times, and application rates used at a particular facility.
Although some of the volatile organics (and nonvolatile organics) may be released in
the form of true aerosols, most would be expected to be released as gases by
evaporation (volatilization). The process of air stripping has been shown to be
particularly efficient in releasing volatile organics at aeration basins of activated
sludge wastewater treatment plants, e.g., the compounds hexachlorobicyclohepta-
diene, heptachlorobicycloheptene, and chlordene at Memphis, Tennessee (Clark el
al. 198 I). Continuous chronic exposures with intermittent acute exposures to these
toxic organics in Memphis may have resulted in significant health effects in the
treatment plant workers.
A study at an overland flow site has suggested that volatilization reduces the levels
of toxic volatile organics in wastewater by 80-100%-, depending on the specific-
substance and application rate (Jenkins el al. 1981). The removal was adequately
described by first-order kinetics. The results indicated that removal by sorption on
suspended matter and sedimentation was not significant.
A certain amount of evaporation occurs duringthe act of spraying itself. Thus, at a
prototype slow rate land treatment system Jenkins and Palazzo (1981) found a 46%
to 69%. removal of eight volatile organics during sprinkler application.
At the M uskegon slow-rate land treatment system, the behavior of four ofthe most
common toxic organics was studied (Clark el al. 1981): trichloroethane, trichloro-
ethylene, tetrachloroethylene, and chloroform (trichloromethane). Air stripping in
the aerated lagoons was shown to be significant for these com pounds. The maximum
concentrations in air, together with the associated wastewater concentrations, of
these compounds immediately downwind ofthe aerated lagoons and spray irrigation
rigs are shown in Table 2. All of these air concentrations are well below the 8-hour
occupational standards of 45.000/ug/ m3, 535,000/Kg; m3, 670,000pg; m3. and 50,000
fjg m3, respectively (ACG1H 1979).
Comparison ofthe wastewater concentrations in Table 2 with the maximum values
found in Table I suggests that an increase of three orders of magnitude over the
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Table 2. Air and Wastewater Concentrations at Muskegon Land Treatment
System
Aerated Lagoons
Spray Irrigation Rigs
Wastewater (/xj/l) Air (/ug/m3) Wastewater (pg/l) Air ijjg/m3)
Trichloroethane
Trichloroethylene
Tetrachloroethylene
Chloroform
ND*
118
8.9
480
90
73
46
202
3
68
40
2.7
9.3
8.6
ND
*ND = Not Detected «0.1 /ug/l).
Muskegon values is the maximum that would ever be likely to occur. Even such an
increase would probably still result in acceptable air concentrations.
Soil and Groundwater
Organic compounds in wastewater may be volatilized, immobilized by adsorption,
or transported through the soil column, possibly to reach the groundwater.
Adsorbed organics may be subsequently chemically or photochemically degraded,
microbially decomposed, or desorbed. A considerable body of research has been
performed on the behavior of pesticides in soil. This research has shown that the
affinity of soil materials for pesticides, and presumably for organics in general,
decreases in the following order (Chang and Page I978):
Organic Matter
Vermiculite
Montmorillonite
Illite
Chlorite
Kaolinite
Iron and aluminum oxides also adsorb organics. Adsorption of organic pesticides
tends to increase with the concentration of functional groups such as amine, amide,
carboxyl, and phenol. Both laboratory and field experiments suggest that, because of
adsorption by soil particles, most pesticide residues remain in surface soils during
land treatment (Chang and Page 1978).
It has recently been shown that for polynuclear aromatic hydrocarbons adsorption
increases with increasing organic carbon content of the soils and increasing "effective
chain length" of the molecule (Means el a/. 1980). The behavior of poly chlorinated
biphenyls (PCBs) in soil has been comprehensively reviewed by Griffin and Chian
(1980), who concluded that PCBs are strongly adsorbed by soil, and that the nature
of the surface, the soil organic matter content, and the chlorine content and/or
hydrophobicity of the individual PCB isomers are factors affecting adsorption.
Adsorption increases with increasing organic matter content of the soil, with
increasing chlorine content, and with increasing hydrophobicity. One study of PCB
percolation through soil columns showed that less than 0.05% of one isomer was
leached in the worst case.
Once organics are immobilized by adsorption on the surfaces of soil particles,
microbial decomposition, or biodegradation, is probably the major mechanism of
their breakdown. Although there are several abiotic mechanisms for chemical
change, nonenzymatic reactions rarely result in appreciable changes in chemical
structure, and it is biodegradation that brings about major alterations and
mineralization of organics (Alexander 1981). The chief agents of this metabolism are
the indigenous heterotrophic bacteria and fungi.
The potential of microbial decomposition for removal of organics is demonstrated
by the experience at two rapid infiltration and one overland flow land treatment sites.
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At Flushing Meadows in Arizona secondary effluent has resulted in no accumulation
of organic carbon in the soil after ten years of operation and 754 m of total infiltration
(Bouwer and Rice 1978). Secondly, the Lake George Village Sewage Treatment Plant
in New York has been applying unchlorinated secondary effluent to natural delta
sand beds by rapid infiltration since 1939(Aulenbachand Clesceri 1978). After about
forty years of daily infiltration rates of 0.08 to 0.30 m/day, there were no indications
that the soil's capacity to treat the applied effluent was approaching exhaustion. The
greatest removal of constituents occurred in the top 10 m of the sand beds. At a
prototype overland flow land treatment system, at the U.S. Army Cold Regions
Research and Engineering Laboratory in New Hampshire, greater than 94% removal
of each of 13 trace organics by volatilization and adsorption was observed (Jenkins el
al. 1983), with removal efficiencies decreasing as application rates increased and
temperature decreased. With the possible exception of PCB, biodegradation resulted
in the absence of contaminant buildup in the surface soil.
Although complete mineralization and detoxication of organic compounds is
common, many compounds are acted on by microorganisms in soils even though
these microorganisms are unable to use them as their sources of nutrient or energy.
The microorganisms are probably utilizing another substrate while performing the
transformations known as "cometabolism" (Alexander 1981). Cometabolism may
lead to detoxication, the formation of new toxic substances, or the synthesis of
persistent products. There is evidence that cometabolism may be particularly
common for toxic organics in very low concentrations in the environment (Rubin el
al. 1982, Subba-Rao el al. 1982).
The metabolism of few chemicals has been studied in microbial cultures, and even
fewer in natural ecosystems. Why certain intermediate compounds in a metabolic
sequence accumulate outside or inside the active organism is not known, and it is
extremely difficult to predict the chemical fate of toxic organics in the environment.
The prediction of biodegradability from chemical structure, although theoretically
possible, has thus far proven problematic. Alexander (1981) has described several
common types of reactions that may occur, and these are listed in Table 3. It used to
be thought that every organic compound could be completely decomposed by
microorganisms. Thus a recent evaluation (Kobayashi and Rittmann 1982) indicated
that the use of properly selected populations of microorganisms, and the
maintenance of appropriate controlled environmental conditions, could be an
important means of improving biological treatment of organic wastes, and that
members of almost every class of synthetic compound can be degraded by some
Table 3. Common Types of Chemical Transformations in the
Environment
Dehalogenation Nitro metabolism
Deamination Oxime metabolism
Decarboxylation Nitrile/amide metabolism
Methyl oxidation Cleavage reactions (many types)
Hydroxylation and Nondegradative reactions:
ketone formation
ft oxidation Methylation
Epoxide formation Ether formation
Nitrogen oxidation N-Acylation
Sulfur oxidation Nitration
=S to =O N-Nitrosation
Sulfoxide reduction Dimerization
Reduction of triple bond Nitrogen heterocycle
Reduction of double bond formation
Hydration of double bond Oligomer and polymer formation
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microorganism. However, field evidence has resulted in the conclusion that some
synthetic organics are decomposed slowly, if at all, and may persist for long periods
in the environment. Alexander (1981) has summarized the possible reasons for this.
A general idea of the relative degree of biodegradation of toxic organics to be
expected in soil may be gained from Tabak el o/.'s (1981) studies on organic priority
pollutants. They collected data on the biodegradability and rate of microbial
acclimation of 96 compounds (5 and 10 mg/1) in a static culture flask screening
procedure, using domestic wastewater inoculum and synthetic medium. Microbial
acclimation, or adaptation, was measured by making three weekly subcultures;
percentage biodegradation was measured after seven days incubation in the dark at
25°C. Their overall results are summarized in Table 4. Significant biodegradation
was found for phenolic compounds, phthalate esters, naphthalenes, and nitrogenous
organics; variable results were found for monocyclic aromatics, polycyclic
aromatics, polychlorinated biphenyls, halogenated ethers, and halogenated ali-
phatics; and no significant biodegradation was found for organochlorine pesticides.
Extrapolation of the above results to the behavior of toxic organics in the soil must
be done with two provisos: (1) biodegradation in soil may be somewhat different
from that in the aquatic medium used for the tests, and (2) the lower concentration of
the organics at the land treatment site may not elicit microbial activity or enzyme
induction. Nevertheless, a comparison of the results with the compounds to be
expected in wastewater, as listed in Table 1, is instructive. Among the top ten
compounds in the table, nine have significant degradation, and one has slow to
moderate degradation with significant volatilization. Among the next ten com-
pounds, eight have significant degradation, two of which are followed by toxicity.
Only two compounds are not significantly degraded, the pesticides heptachlor and
lindane.
Other than for pesticides, the literature on the microbial decomposition of toxic
organics in soil is sparse. The degradation of petroleum hydrocarbons, a mixture of
aliphatic, aromatic, and asphaltic compounds, has been reviewed by Atlas (1980,
1981). Factors which appear to be important in encouraging high decomposition rate
of petroleum hydrocarbons are high temperature, low concentrations, high soil
fertility, and an aerobic environment. There is little evidence for significant
downward leaching of oil. Experiments with the high-rate application of high
petroleum hydrocarbon sludge to land have shown a 77% degradation rate near the
surface after one year, most of the degraded compounds being n-alkanes (Lin 1980),
and it was concluded that sludge land disposal would not result in petroleum
hydrocarbon buildup in the soil. In studies of organic substances in wastewaters used
for irrigation, Dodolina el al. (1976) found acetaldehyde, crotonaldehyde, benzal-
dehyde, cyclohexanone, cyclohexanol, and dichloroethane to disappear from soil
within ten days. The biodegradation in soil of polychlorinated biphenyls was
reviewed by Griffin and Chian (1980) who concluded that they are degradable, but
that resistance increases as isomers have higher chlorination. Polybrominated
biphenyls, on the other hand, have shown little biodegradation after one year in soil
(Jacobs et al. 1978).
In view of the multitudinous variety of organic compounds existing, it is difficult to
generalize about their biodegradation in soil. It appears, however, that most organics
do become microbially decomposed in the soil, at least to some extent. This is
especially true of naturally-occurring compounds, or those resembling them, because
of the eons of evolution that have developed microbial enzyme systems to do thejob.
The more structually complex the molecule is, e.g., condensed rings or dense
branching, and more halogenated it is, the more difficult is biodegradation. A
convenient sampling of the international literature on decomposition of organics in
soil may be found in Overcash (1981).
It has recently been found that it might be possible for carcinogenic and
teratogenic nitrosamines to be formed from secondary and tertiary amines at land
treatment sites. Thus, Thomas and Alexander (1981) have shown that dimethylamine
10
-------
Table 4. Biodegradability of Priority Organic Compounds (after Tabak et a/. 1981)
D—Significant degradation with rapid adaptation
A—Significant degradation with gradual adaptation
T—Significant degradation with gradual adaptation followed by a deadaptive process (toxicity)
B—Slow to moderate biodegradative activity, concomitant with significant rate of volatilization
C—Very slow biodegradative activity, with long adaptation period needed
N—Not significantly degraded under the conditions of test method
Test Compound
Performance
Summary
Test Compound
Performance
Summary
Phenols
Phenol
2-Chloro phenol
2,4-Dichloro phenol
2,4,6-Trichloro phenol
Pentachloro phenol
2,4 Dimethyl phenol
Phthalate Esters
Dimethyl phthalate
Diethyl phthalate
Di-n-butyl phthalate
Naphthalenes
Naphthalene
2-Chloro naphthalene
Acenaphthene
Acenaphthylene
D
D
D
D
A
D
D
D
D
D
D
D
D
p-Chloro-m-cresol
2-Nitro phenol
4-Nitro phenol
2,4-Dinitro phenol
4,6 Dinitro-o-cresol
Bis-(2-ethyl hexyl) phthalate
Di-n-octyl phthalate
Butyl benzyl phthalate
D
D
D
D
N
A
A
D
-------
Table 4. (Continued)
D—Significant degradation with rapid adaptation
A—Significant degradation with gradual adaptation
T—Significant degradation with gradual adaptation followed by a deadaptive process (toxicity)
B—Slow to moderate biodegradative activity, concomitant with significant rate of volatilization
C—Very slow biodegradative activity, with long adaptation period needed
N—Not significantly degraded under the conditions of test method
Test Compound
Monocyclic Aromatics
Benzene
Chlorobenzene
1,2-Dichlorobenzene
1,3-Dichlorobenzene
1,4-Dichlorobenzene
1,2,4-Trichlorobenzene
Poly cyclic Aromatics fPAHsJ
Anthracene
Phenanthrene
Fluorene
Fluoranthene
Po/ych/orinated Biphenyls (PCBs)
Aroclor-1016
Aroclor-1221
Aroclor-1232
Aroclor-1242
Performance
Summary
D
D-A
T
T
T
T
A
D
A
A-N
N
D
D
N
Test Compound
Hexachlorobenzene
Nitrobenzene
Ethylbenzene
Toluene
2,4-Dinitrotoluene
2,6-Dinitrotoluene
1,2-Benzanthracene
Pyrene
Chrysene
Aroclor-1248
Aroclor-1254
Aroclor-1260
Performance
Summary
N
D
D-A
D
T
T
N
D-N
A-N
N
N
N
-------
Table 4. (Continued)
D—Significant degradation with rapid adaptation
A—Significant degradation with gradual adaptation
T—Significant degradation with gradual adaptation followed by a deadaptive process (toxicity)
B—Slow to moderate biodegradative activity, concomitant with significant rate of volatilization
C—Very slow biodegradative activity, with long adaptation period needed
N—Not significantly degraded under the conditions of test method
Test Compound
Performance
Summary
Test Compound
Performance
Summary
Ha/ogenated Ethers
Bis-(2-chloroethyl) ether
2-Chloroethyl vinyl ether
4-Chlorodiphenyl ether
Nitrogenous Organics
Nitrosamines
N-Nitroso-di-N-propylamine
N-Nitrosodiphenylamine
Substituted benzenes
Isophorone
1,2-Diphenylhydrazine
Ha/ogenated A/iphatics
Chloroethanes
1,1 -Dichloroethane
1,2-Dichloroethane
1,1,1 -Trichloroethane
1,1,2-Trichloroethane
1,1,2,2-Tetrachloroethane
D
D
N
N
D-A
D
T
A
B
B
C
N
4-Bromodiphenyl ether
Bis-(2-chloroethoxy) methane
Bis-(2-chloroisopropyl) ether
Acrylonitrile
Acrolein
Chloroethylenes
1,1 -Dichloroethylene
1,2-Dichloroethylene-cis
1,2-Dichloroethylene-trans
Trichloroethylene
Tetrachloroethylene
N
N
D
D
D
A
B
B
A
A
-------
Table 4. (Continued)
D—Significant degradation with rapid adaptation
A—Significant degradation with gradual adaptation
T—Significant degradation with gradual adaptation followed by a deadaptive process (toxicity)
B—Slow to moderate biodegradative activity, concomitant with significant rate of volatilization
C—Very slow biodegradative activity, with long adaptation period needed
N—Not significantly degraded under the conditions of test method
Test Compound
Hexachloroethane
Halomethanes
Methylene chloride
Bromochloromethane
Carbon tetrachloride
Chloroform
Dichlorobromomethane
Bromoform
Chlorodibromomethane
Trichlorofluoromethane
Organochlorine Pesticides
Aldrin
Dieldrin
Chlordane
DDT p,p'
DDE p,p'
ODD p,p'
Endosulfan-alpha
Endosulfan-beta
Endosulfan sulfate
Endrin
Performance
Summary
D
D
D
D
A
A
A
N
N
N
N
N
N
N
N
N
N
N
N
Test Compound
Chloropropanes
1,2-Dichloropropane
Chloropropylenes
1,3-Dichloropropylene
Chlorobutadienes
Hexachloro-1,3-butadiene
Chloropentadienes
Hexachlorocyclopentadiene
Heptachlor
Heptachlor epoxide
Hexachlorocyclohexane
a BHC-alpha
Hexachlorocyclohexane
P BHC-beta
Hexachlorocyclohexane
<5 BHC-delta
Hexachlorocyclohexane
y BHC-gamma (lindane)
Performance
Summary
A
A
D
D
N
N
N
N
N
N
-------
and trimethylaminc can be formed in municipal wastewater from naturally-
occurring precursors. Dimethylamine may then go on to be microbially nitrosated,
forming N-nitrosodimethylamine, a process which can occur under conditions
resembling land treatment of wastewater (Greene el at. 1981). Whether this can
actually occur under field conditions, resulting in a threat to groundwater, has not
been demonstrated.
The transport to groundwater of organics has been studied at only a few operating
land treatment sites. Bouwer and Rice (1978) have noted that the total organic
carbon concentration of the renovated water at the Flushing Meadows rapid
infiltration site, in Phoenix, Arizona, averaged a bout 4 mg/1, indicating the presence
of refractory trace organics. They felt that these organics were the main health
concern for possible potable use of the groundwater, since their identity is not
completely known and carcinogenicity or other toxicity is suspected. More detailed
data on groundwater contamination by organics is now available from the nearby
23rd Avenue Project rapid infiltration site, also in Phoenix, Arizona (Tomson el al.
1981). At this site secondary effluent is treated by about 25 feet of soil. The overall
removal rate of refractory volatile organics (R VO) at this site was 86.4%, resulting in
a ground water concent rat ion of RVO of 2.245/vg/1. Removal rates are presented, by
class, in Table 5. The most likely volatile organic contaminants of groundwater at
this site are the chloroalkanes, alkylphenols, alkanes, and amides.
TableB. Removal Efficiencies
of Refractory Volatile
Organics at 23rd Ave-
nue Project Rapid In-
filtration Site (after
Tomson et al. 1981)
Class
Chloroalkanes
Chloroaromatics
Alkylbenzenes
Alkylphenols
Alkylnaphthalenes
Alkanes
Alcohols
Ketones
Indoles, Indenes
Amides
Alkoxyaromatics
70
94
98
85
100
71
95
98
96
74
91
Data for the transport of organics to groundwater exist from two slow-rate land
treatment sites, a prototype system operated by the U. S. Army Cold Regions
Research and Engineering Laboratory in New Hampshire (Jenkins and Palazzo
1981) and the Muskegon system in Michigan. The New Hampshire prototype site
showed good removals for natural chloroform and toluene in the wastewater, and
greater than 98% removal for four spiked volatile organics (Table 6). At the
Muskegon system (Demirjian et al. 1983) the irrigation water contained 1-2 /ug/1
levels of chloroform, 1,2-dichloroethane, 2-chloroaniline, and 2,2'-dichloroazo-
benzene after lagoon pretreatment. Only two organic compounds were detected as
residuals in the soil: xylene, from the dispersing agent used with herbicides, and
2,2-dichloroazobenzene, less than 0.15 mg/kg and possibly derived from photo-
chemical conversion of 2-chloroaniline. There was no evidence for accumulation of
15
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Table 6. Removal of Volatile Organics at a Prototype Slow Rate Land
Treatment Site (after Jenkins and Palazzo 1981)
Mean Concentration (/ug/l)
Wastewater
Substance Before Spraying
Chloroform
Toluene
Methylene chloride
1,1 -Dichloroethane
Bromodichloromethane
Tetrachloroethylene
41.8
57.3
7.61
30.2
11.1
61.9
Wastewater
After Spraying
14.0
24.4
2.32
9.88
3.98
22.7
Test Cell 1
Percolates
0.86
0.06
0.06
b.d.*
b.d.*
0.08
Test Cell 6
Percolates
0.73
0.02
0.04
0.06
0.02
0.35
*Below a detection limit of about 0.01 fjg/\.
other organics after seven years of operation. In the drain tile effluent (at 5-12 feet
depth) there was sporadic occurrence of approximately 1 /jg/1 levels of chloroform,
1,2-dichloroethane, and 2-chloroaniline, the latter probably from lagoon seepage.
Also in the effluent were 1 /vg/1 levels of the herbicides simazine and atrazine.
From these data it is evident that toxic organics can be transported to groundwater
below land treatment sites, although the degree can be controlled by the level of
preapplication treatment and application rate as well as choice of effluent and site
characteristics. This certainly is cause for concern, but it should be kept in mind that
groundwater is not the pristine substance it was once thought to be(Burmaster 1982).
The synthetic organic compounds most commonly found in groundwater in the U.
S., deriving primarily from industrial wastes, are (Environmental Health Letter
2!(6):7, 1982):
trichloroethylene benzene
tetrachloroethylene chlorobenzene
carbon tetrachloride dichlorobenzene
1,1,1 -trichloroethane trichlorobenzene
1,2-dichloroethane 1,1 -dichloroethylene
vinyl chloride cis-l,2-dichloroethylene
methylene chloride trans-1,2-dichloroethylene
Once organic compounds gain access to groundwater, they may be subject to some
of the same removal processes that affect them at the surface, particularly adsorption
and microbial decomposition, although certainly at much lower rates. These two
processes, which largely govern the movement and fate of organics in the subsurface
environment, have been reviewed by McCarty et a/.(1980, 1981). The degree of
adsorption of an organic compound in groundwater is to a great extent dependent
upon its hydrophobicity, especially when the aquifer organic content is above about
0.1 %. Thus, only compounds with octanol/ water partition coefficients less than 103
are likely to readily move through the subsurface environment. Of course, these are
the very compounds most likely to reach the groundwater at a land treatment site, the
more hydrophobic compounds having been adsorbed to the soil above. Likewise, it is
probable that most microbial decomposition would have occurred before the
organics reach the groundwater, although there is evidence that diverse microbial
populations of sulfate reducers, methanogens, and heterotrophs exist and are
metabolically active in aquifers, and that biodegradation of some organic pollutants
occurs in groundwater (Gerba and McNabb 1981). Nevertheless, it is difficult to
avoid the conclusion that once toxic organics get into the groundwater they will
remain there for a long time.
16
-------
Plants
At the low concentrations found in the soil at municipal wastewater land treatment
sites, very few organic compounds are likely to be toxic to plants. In a review of data
on over 130,000 chemicals, Kenaga (1981) found only 0.17% of the chemicals killed
seeds or seedlings at concentrations of 0.1-0.99 ppm. Crop plants, however, although
not injured themselves, may accumulate organics that may be toxic to the animals to
which they are fed or to humans who use them as food, either directly or through
animal products.
Among the organics, the pesticides appear to be the most notorious accumulators
in crop plants. Thus, heptachlor, dieldrin, and chlordane are absorbed at low levels
from the soil (Braude et al. 1978). Most herbicides, of course, are readily taken up and
translocated within plants, but there is no reason to think that herbicides would
present more of a problem at land treatment sites than they do at ordinary
agricultural sites.
In contrast with pesticides, most organic compounds are only poorly absorbed and
translocated by plants, with much of the "absorption" probably accounted for by
root adsorption. The literature, however, in this area is sparse. Irrigation of
vegetables in test plots with contaminated wastewaters has shown no accumulation
of polycyclic aromatic hydrocarbons, especially benzo(a)pyrene (Il'nitskii et al.
1974). Trace levels of polychlorinated biphenyls (PCBs) from municipal sludge
applied to an old field has resulted in no detectable PCBs in plant samples (Davis et
al. 1981). Higher levels (50-100 ppm dry soil) resulted in 3-50% of the soil
concentration in carrots (Iwata et al. 1974), with concentration increasing with lesser
chlorinated biphenyls. Since 97% of the PCB was found in the carrot peel, very little
translocation occurred in the plant tissue. On the basis of greenhouse and field
studies of polybrominated biphenyls (PBBs), it has been concluded that little, if any,
PBB will be translocated from contaminated soil to plant tops, and although some
root crops from highly contaminated soil might contain traces of PBB, much of this
PBB could probably be removed by peeling (Chou el al. 1978). 4-Chloroaniline and
3,4-dichloroaniline can be absorbed by tomato plants, oats, barley, and wheat, but
90-95% remains in the roots (Fuchsbichler el al. 1978); in carrots, however, the
chloroanilines are translocated to the upper parts of the plants in significant
quantities. In a study of aldehydes and other organics at agricultural land treatment
sites Dodelina et al. (1976) found no uptake of acetaldehyde, crotonaldehyde, and
benzaldehyde in the aboveground portions of potatoes and corn. Cyclohexanone
and cyclohexanol could be found in corn plants four days after irrigation, but not
later. Dichloroethane was taken up by beets and cereals, but was metabolized and
absent within about two weeks after irrigation.
At the operating land treatment site in Muskegon, corn crop samples for 1980 did
not contain detectable levels of any of the chemicals tested, and it was concluded that
plant uptake of irrigated organic chemicals does not occur to any measurable extent
(Demirjian et al. 1983).
Animals
The low levels of toxic organics to be expected in the aboveground portions of
plants growing at land treatment sites probably pose little hazard to animals feeding
upon them. Under certain site-specific conditions, however, high concentrations of
particular organics in the wastewater may cause problems. For example, PCBs in
cabbage grown on sludge-amended soil have probably caused degenerative changes
in liver and thyroid of sheep (Kienholz 1980), and one can extrapolate a similar
phenomenon to a land treatment site.
A more serious route of exposure by animals to toxic organics is the soil itself.
M ost grazing animals ingest a certain amount of soil together with their food plants.
Thus, dairy cows may ingest 100-500 kg of soil per year, with an average of about
17
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200-300 kg/yr; expressed in other terms, dairy cows may consume soil up to 14% of
dry matter intake when available forage is low and no supplemental feed is used
(Kienholz 1980, Fries 1980). These conditions, however, would probably not be
characteristic of a land treatment site. Lipophilic organics present in the soil may
concentrate in animal fat. For example, feeding experiments with PCBs indicate that
the steady-state milk fat concentrations are about five times the diet concentrations,
which could result in milk fat levels of 0.7 ppm for each 1 ppm of PCBs in surface soil
(Fries 1980). Body fat levels would be expected to be similar.
Conclusions and Research Needs
The tremendous number of organic compounds possibly present in wastewater,
together with their myriad health effects and poorly understood behavior in the
environment, represent a considerable potential for adverse health effects. Most of
these can probably be prevented by simple design and monitoring measures; this, of
course, would not be true in the case of high discharges of particular chemicals.
Preapplication treatment by storage lagoons may remove considerable quantities
of organics, but cannot be relied upon to efficiently remove all toxic organics,
particularly since most pretreatment design questions center on inactivation of
pathogens.
Although removal rates of organics from wastewater by aerosolizeation and
volatilization are high, exposure through this route is unlikely to present any
significant health effect.
Toxic organics can enter the groundwater, particularly at rapid infiltration sites,
and the application and soil factors controlling this transport, together with the
factors governing their movement and decomposition within groundwater, are
significant research needs.
The levels of toxic organics likely to be present in soils at land treatment sites will
probably result in extremely low levels in above-ground portions of plants, but levels
in roots, tubers, and bulbs may present a health hazard. The feeding of land-
treatment-site-grown plants to animals is unlikely to pose a health problem, but
grazing animals may accumulate significant levels of toxic organics. The issue of
accumulation of organics from the soil by plants and animals (particularly into milk)
is poorly understood, and more research is required.
18
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TRACE ELEMENTS
Types and Levels in Wastewater
The trace elements (including the "heavy metals") in wastewater of public health
concern, i.e., those for which primary drinking water standards (USEPA 1977) exist
(but excluding silver since its effect is largely cosmetic), are:
Primary Drinking Water
Standard (mg/1)
Arsenic (As) 0.05
Barium (Ba) 1.0
Cadmium (Cd) 0.010
Chromium (Cr) 0.05
Lead(Pb) 0.05
Mercury (Hg) 0.002
Selenium (Se) 0.01
Of these, cadmium, lead, and mercury are usually regarded as of most concern, and
barium of minor concern. Chromium and selenium are essential elements in man;
arsenic and cadmium have been shown to be essential to experimental animals and,
thus, may be essential to man as well (National Research Council 1980). Secondary
drinking water standards (USEPA 1979), i.e., those related to aesthetic quality, also
exist for copper, iron, manganese, and zinc. These latter elements, as well as all other
trace elements, are toxic if ingested or inhaled at high levels for long periods
(Underwood 1977), but this fact does not warrant considering them in the land
treatment context, where low levels are expected.
Arsenic is popularly known as an acute poison, but chronic human exposure to low
doses, as might be expected for all trace elements as a result of land treatment, may
cause weakness, prostration, muscular aching, skin and mucosal changes, peripheral
neuropathy, and linear pigmentations in the fingernails. Chronic arsenic intoxication
may result in headache, drowsiness, confusion, and convulsions (Underwood 1977).
Epidemiological evidence has implicated arsenic as a carcinogen, but there is little
evidence that arsenic compounds are carcinogenic in experimental animals
(Sunderman 1977). Even with high concentrations in soil, however, plants rarely take
up enough of the element to constitute a risk to human health (Underwood 1977,
Council for Agricultural Science and Technology 1976).
Barium has a low degree of toxicity by the oral route. Because of its effect of
intensely stimulating smooth, striated, and cardiac muscle in acute exposures,
however, it may have cardiovascular effects in low doses, but this has not thus far
been demonstrated (Brenniman et al. 1979).
Cadmium is widely regarded as the trace element of most concern from a human
health effects viewpoint in the land application of sludge, and this status probably
carries over into the land treatment of wastewater as well. The critical health effect of
chronic environmental exposure via ingestion is renal tubular damage due to
accumulation of cadmium in the kidney. The initial consequence of this damage is the
loss of low molecular weight serum proteins in the urine, followed by loss of other
proteins, glucose, amino acids, and phosphate. This kidney damage is often
irreversible and constitutes a significant adverse health effect (Ryan et al. 1982).
There is evidence that the absorption and/or toxicity of cadmium are antagonized by
19
-------
zinc, selenium, iron, and calcium (Sandstead 1977). The carcinogenicity of cadmium
is controversial; the epidemiological evidence is tenuous, and the experimental
evidence is conflicting (Ryan et al 1982).
Chromium is much more toxic in its hexavalent form than its trivalent form, its
predominant state in wastewater and soil. Chronic oral exposure in experimental
animals has been associated with growth depression, and liver and kidney damage
(Underwood 1977). Hexavalent chromium causes respiratory cancer upon chronic
exposure to chromate dust (Sunderman 1977). Most crops absorb relatively little
chromium from the soil (Council for Agricultural Science and Technology 1976).
Lead chronic toxicity is characterized by neurological defects, renal tubular
dysfunction, and anemia. Damage to the central nervous system is common,
especially in children, who have low lead tolerance, resulting in physical brain
damage, behavioral problems, intellectual impairment, and hyperactivity. At soil pH
above 5.5 and high labile phosphorus content, common conditions at a land
treatment site, little movement of lead from the soil into plant tops and seed would be
expected (Council for Agricultural Science and Technology 1976, Stewart 1979).
Mercury in low levels can result in neurological symptoms such as tremors,
vertigo, irritability, and depression, as well as salivation, stomatitis, and diarrhea.
Mercury can enter plants through the roots, and appears to be readily translocated
throughout the plant (Council for Agricultural Science and Technology 1976),
although there is some contrary evidence (Stewart 1979).
Selenium exposure in its chronic form is associated with dental caries, jaundice,
skin irruptions, chronic arthritis, diseased finger and toenails, and subcutaneous
edema. It has also been found to have an inhibitory effect against several types of
cancer (Fishbein 1977). Selenium is readily taken up by plants and passed on to
animals, and has caused toxicity in livestock in high-selenium soils (Council for
Agricultural Science and Technology 1976, Underwood 1977).
Ranges of levels of trace elements in untreated municipal wastewater in the United
States are presented in Table 7, together with recommended irrigation water quality
criteria for comparison. It is evident that the trace elements in wastewater most likely
to violate agricultural irrigation criteria are cadmium and chromium. The high levels
of these two elements found in certain municipal wastewaters are doubtlessly due to
industrial sources, for example electroplating operations.
Since chromium is poorly absorbed from the soil by crops, cadmium is probably
the element of most public health concern. That the upper limits of the cadmium
ranges appearing in Table 7, i.e., 0.14 and 1.80 mg/1, might be unusual for municipal
wastewaters is suggested by recent data for Chicago (Lue-Hing et al. 1980), a highly
industrialized city. In 1977 five wastewater treatment plants in the Metropolitan
Sanitary District of Greater Chicago had average raw wastewater cadmium
concentrations of 0.045, 0.021, 0.005, 0.011, and 0.018 mg/1, all below the less
restrictive 20-year irrigation criterion of 0.050 mg/1. Purely domestic wastewater,
i.e., with no industrial input, in the Chicago area has cadmium concentrations of
0.0011-0.0022 mg/1, of which 0.0002-0.0013 mg/1 is due to the local tap water
(Gurnham et al 1979).
Preapplication Treatment
The removal of trace elements from wastewater by primary treatment (screen, grit
chamber, and sedimentation tank) has been reported (Crites and Uiga 1979) to be
approximately:
Cadmium 30%
Chromium 40%
Lead 50%
One would expect wastewater stabilization ponds to achieve at least as high removal
rates.
20
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Table?. Concentration of Trace Elements in Untreated Municipal Waste
water in U.S. and Recommended Irrigation Water Criteria (mg/l)
Untreated Wastewater Irrigation Criteria1
Element
Arsenic
Barium
Cadmium
Chromium
Lead
Mercury
Selenium
USEPA 1981
0.003
0.004-0.14
0.02-0.700
0.05-1.27
0.002-0.044
—
Feiler 1980
0.002-0.080
0.001-1.800
0.008-2.380
0.016-0.935
0.0002-0.0039
0.001 -0.020
All Soils2 Fine-Textured Soils3
0.10
0.010
0.10
5.0
—
0.020
2.0
0.050
1.0
10.0
—
0.020
'NAS-NAE 1972.
2For waters used continuously on all soils.
3For use up to 20 years on fine-textured soils of pH 6.0 to 8.5.
Soil and Plants
Most of the trace element content of wastewaters appears to be associated with
finely-divided suspended solids (Brown 1978, Chang and Page 1978). These are
removed near the soil surface by straining and filtration. The removal of the soluble
portion depends upon the texture, clay, and organic matter content of the soil, as well
as the chemical properties of the specific element, with most soils possessing the
capacity to immobilize trace elements near the surface by precipitation reactions and
adsorption. Cation exchange reactions do occur, but are not expected to play a major
role because of the concentration effect in competition for exchange sites. In any
case, cation exchange results in temporary removal from the soil solution.
Precipitation reactions include the formation of poorly-soluble oxides, hydroxides,
carbonates, phosphates, sulfites, etc., for the cations, and formation of anions for
arsenic and selenium. Mercury, of course, may leave the soil through volatilization.
The soil chemistry of most of the individual toxic trace elements has been concisely
summarized by Chang and Page (1978).
Limits for the maximum cumulative application of trace elements to agricultural
land have been recommended by various governmental agencies, for the protection
of public health and the prevention of phytotoxicity. These have almost invariably
been proposed in the context of the land application of sludge, but should be just as
valid for land treatment of wastewater, at least in the slow rate mode. Limits for rapid
infiltration and overland flow could be less restrictive because of the lack of
production of crops for human consumption and the greater depth of soil involved in
treatment (in the former case). These limits have been used, together with typical
wastewater levels of trace elements, by Page and Chang (1981) to predict the useful
life of a typical land treatment site where crops are grown for human consumption.
The results appear in Table 8. It is evident that cadmium, with a 17-67 year limit, is
the element most likely to restrict the use of wastewater for irrigation of crops for
human consumption. In the case of crops not for human consumption, other
elements may be limiting—in particular molybdenum because of its toxicity to
livestock, and nickel because of its phytotoxicity. These latter elements yield limits of
47-48 years in Page and Chang's analysis.
Trace elements, of course, are conservative materials, in contrast to organics and
pathogens, which may become inactivated and decomposed. Thus, one would expect
cadmium to build up in the soil at a land treatment or irrigation site. Our major hope,
from a health effects point of view, is that it will become immobilized in the soil, and
less available for plant uptake as time passes. There is some suggestion of this in
Hinesly's data (Hinesly et al 1979) on cadmium content of corn grain and leaves,
grown during and after the termination of wastewater sludge application (Table 9). It
is of interest to note that the cadmium content of the grain approached that of control
21
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Tables. Annual Input of Trace Elements and Years of Land Treatment
Required to Exceed Recommended Cumulative Input Limits
(modified from Page and Chang 1981)
Element
Arsenic
Cadmium
Chromium
Lead
Mercury
Selenium
Wastewater (mg/l)
0.005
0.02
0.05
0.25
0.0009
0.005
Annual
Input'
(kg/ha)
0.075
0.3
0.75
3.0
0.014
0.075
Recommended
Cumulative
Input Limits
(kg/ha)
USA2 UK3
— 10
5/10/20" 5
— 1000
800 1 000
2
— 5
Years Required to
Exceed Limits
USA UK
— 133
17/33/67" 17
— 1333
267 333
— 143
— 67
'Assuming an annual application rate of 1.5 m.
2USEPA, USFDA, and USDA 1981.
National Water Council 1977.
"For soils with cation exchange capacities of <5, 5-15, and >15 meq/100g, respec-
tively, and soil pH >6.5. If soil pH <6.5, first figure holds.
5Raised over the value presented in Page and Chang (1981), to reflect the data in
Table 7.
grain within three years after termination of sludge applications, even though the
cumulative cadmium application was 58.3 kg/ha, almost three times the maximum
recommended limit of 20 kg/ ha. The issue of cadmium uptake by plants after annual
vs. cumulative application limits and the effect of discontinuing application remain
unsettled, however, (Ryan et al 1982).
Examination of trace elements in soil and plants at long-term operating municipal
land treatment sites has yielded mixed results. Three locations in the U. S. where
crops (in two sites) have been irrigated for 17-33 years showed little or no
accumulation of trace elements in soil: Roswell, New Mexico, slow rate site (Koerner
and Haws 1979a), Tooele, Utah, slow rate site (Reynolds et al. 1979), and Vineland,
New Jersey, rapid infiltration site (Koerner and Haws 1979b). After 18 years of
slow-rate irrigation with wastewater at San Angelo, Texas, there was no significant
change in total soil cadmium concentration, but a six-fold increase of total soil lead
concentration (Hassner el al. 1978). Wastewater concentrations were <0.004-0.017
mg/l cadmium and <0.050-0.230 mg/l lead, at the low end of the ranges given in
Table 9. Cadmium Uptake by Corn During and After Wastewater Sludge
Application (after Hinesly et al. 1979)
Year
1969
1970
1971
1972
1973
1974
1975
1976
1977
Sludge
Applied
(t/ha)
16.4
52.8
57.8
44.4
61.1
0
0
0
0
Cd
(kg/ha)
7.9
22.6
13.2
7.8
6.8
0
0
0
0
Cd in Grain (/jg/g)
Control
—
—
0.15
0.15
0.16
0.18
0.15
0.14
0.10
Sludged
—
—
1.37
0.89
0.44
0.23
0.17
0.15
0.07
Cd in Leaves Oug/g)
Control
—
0.5
0.5
0.3
0.1
0.7
0.3
0.3
Sludged
—
—
35.6
23.3
7.1
3.6
5.9
2.9
2.1
22
-------
Table 7. Slightly lower levels of cadmium in wastewater«0.001-0.008 mg/1) led to a
doubling of soil cadmium concentration after 30 years of rapid infiltration at
Hollister, California (Pound et al. 1978). That trace elements can increase in plants at
land treatment sites is documented by results at Werribee Farm in Melbourne,
Australia, where, after over 70 years of primary effluent irrigation trace element
levels in both soil and plants have increased with cumulative input loads (Croxford
1978). Wastewater, soil, and pasture plant levels of toxic trace elements are
summarized in Table 10. Similar levels of cadmium, but lower levels of chromium,
have been found in vegetables grown at a site in Santiago, Chile, where the city's
untreated wastewater has been applied for over 40 years (Schalscha et al. 1978).
Groundwater
Slow-rate land treatment appears to stabilize trace elements, and prevent their
entry into the groundwater. Thus, after a year of application of cadmium, copper,
lead, nickel, and zinc-amended wastewater to soil columns in the laboratory. Brown
el al. (1983) found almost all of the metals to remain within the top 25 cm, and none to
pass into the leachate at 1.5 m depth. At the prototype land treatment system in
Hanover, New Hampshire, trace elements were found to be removed quickly in the
first several centimeters of soil, and did not seem to move deeper into the profile, even
after five years following termination of the trace element spiking into the wastewater
(Jenkins and Palazzo 1981). After 18 years of slow-rate irrigation with wastewater at
the San Angelo site, concentrations of cadmium, chromium, and lead were below
drinking water standards in seepage creeks, shallow ground wells, and deep wells
within the sewage farm (Hossner el al. 1978).
Rapid-infiltration sites, in contrast, appear to present some threat to groundwater.
After 30 years of primary wastewater infiltration at the Hollister site, concentrations
in shallow groundwater of cadmium (0.028 mg/1) and lead (0.09 mg/1) were above
drinking water standards, but arsenic, barium, chromium, mercury, and selenium
were below (Pound et al. 1978).
Animals
The accumulation of trace elements in cattle grazed on sludge-amended pastures
has revealed raised levels in liver and kidney, but not in muscle tissue (Bertrand et al.
1981a). No increases were seen when cattle were fed sludge-amended-soil-grown
forage sorghum (Bertrand et al 1981 b). It may not be fair, however, to extrapolate
from sludge amended land, where high trace element soil levels may be expected, to
land treatment sites. Experience at Werribee Farm in Melbourne, Australia, where
cattle are grazed on wastewater-irrigated pastures, has shown higher organ levels of
cadmium and chromium than in Farm cattle grazed on non-irrigated pastures, but
comparable to non-Farm cattle (Table 11). Organ levels of lead, however, did not
increase, in spite of increases in both soil and pasture plants (see Table 10).
Since trace elements accumulate in very small quantities in animal muscle tissue,
there is probably little concern about non-visceral meats in the marketplace. Liver
and kidneys of animals do, however, accumulate high levels of cadmium, just as they
do in man, so that these meats may be of concern to those people consuming large
quantities of them.
Conclusions and Research Needs
It seems reasonable to conclude that cadmium is the only trace element likely to be
of health concern to humans as a result of land treatment of wastewater, with the
exposure being through food plants or organ- meats. Groundwater is unlikely to
represent a threat except at rapid infiltration sites.
23
-------
Table 10. Wastewater, Soil, and Pasture Plant Levels of Toxic Trace Elements at Werribee Farm (after Croxford
1978)
Soil (/ug/g dry wt)
Trace
Element
Cadmium
Chromium
Lead
Raw
Wastewater
(mg/l)
0.015
0.04
0.03
Normal
(mean &
range)
0.06
(0.01-7)
100
(5-3000)
10
(2-200)
Farm
Control
(0-5 cm)
0.5
13-22
11-29
Farm
Irrigated
(0-5 cm)
0.5-3.9
90-325
56-241
Pasture Plants
(/jg/g dry matter)
Normal
0
0
0.
.2-0.8
.2-1.0
1-10
Farm
Control
0.19
2.3
3.4
Farm
Irrigated
1.1
15
12
-------
Table 11. Toxic Trace Element Concentrations in Cattle Liver and
Kidney at Werribee Farm (fjg/g dried tissue) (after Croxford
1978)
Cadmium Chromium Lead
Cattle Liver
Non-Farm
Farm: Non-Irrigated
Farm: Irrigated
Cattle Kidney
Non-Farm
Farm: Non-Irrigated
Farm: Irrigated
0.76
0.17
0.38
3.32
1.24
2.07
0.05
0.05
0.07
0.06
0.05
0.07
0.5
0.93
1.12
0.32
2.24
1.41
The significance of this concern with cadmium getting into the human food chain
depends upon the cadmium levels presently exist ing in human food, the total dietary
intake of cadmium, and the potential increase in cadmium levels in human food due
to land treatment.
The cadmium levels presently existing in human food can be estimated, at least for
the United States, by data from the U. S. Food and Drug Administration's
Compliance Program ("market-basket survey"). These levels, together with the
calculated normal dietary intake and vegetarian dietary intake of cadmium, are
summarized in Table 12. It should be noted that root and leafy vegetables have the
highest concentrations of cadmium. More accurate estimates of cadmium (and other
trace element) concentrations in crops grown in the U. S., together with
concentrations in the soils in which they are growing, will soon be available from a
survey jointly supported by the USEPA, USFDA, and U. S. Department of
Agriculture. In this survey 6,000 crop samples and 18,000 soil samples are being
analyzed over a four year period, and the results should be available in the near
future.
The present total dietary intake of cadmium was estimated (Table 12) to be about
28 yUg/day. Other estimates based on the market-basket method have resulted in
higher values: 26-61 /ug/day in 15-20-year old U.S. males, by the USFDA, and 52
/ug/day in Canadians (Kirkpatrick and Coffin 1977).
A more direct, and potentially more accurate, method of estimating dietary intake
of cadmium is by measuring the cadmium content of human feces. This method is
feasible because the absorption of cadmium from the gut is low—rarely more the
10%, and usually 4-6%—and the excretion of cadmium into the gut is also very low. It
is more accurate because cadmium concentration is generally about ten times more
concentrated in feces than food, and because feces reflects actual food intake rather
than predicted. A recent study, using existing fecal cadmium data in Chicago and
Dallas, and estimating daily feces production, resulted in a final estimate of the
average daily intake of cadmium in food for U. S. inhabitants of 13-16 fjg/day
(Kowal ei al. 1979). (Since the ingestion rate of the teenage male is often used in
discussions of cadmium intake, values of 24/jg/day, 19/yg/day, and 18 Aig/day for
10-19-year-old males from Chicago 1974, Chicago 1976, and Dallas, respectively,
were estimated.)
This estimate of the average daily intake of cadmium in food can be compared with
other estimates by the fecal analysis method, where the daily feces production of each
individual was measured rather than estimated. In Sweden rates of 16 /Kg/day in
nonsmokers and 19/ug/day in smokers (former and present) have been reported (see
Kowal el al. 1979 for references). The increased rate for smokers was partly
attributed to their increased food intake. Nine /ug/day fecal cadmium has been
25
-------
Table 12. Cadmium Concentration in Foods and Calculated Dietary
Intake (from Ryan era/. 1982)
Normal Diet
Vegetarian Diet0
Food Classes
Dairy products
Meat, fish,
poultry
Grain & cereal
products
Potatoes
Leafy vegetables
Legume vegetables
Root vegetables
Garden fruits
Fruits
Oily fats,
shortenings
Sugars & adjuncts
Beverages
Total Intake
ppb Cda
5.7
15.3
23.2
48.0
40.5
6.2
32.3
14.7
3.0
15.3
10.0
3.0
g/day
549
204
331
138
42
51
25
69
173
56
65
534
2,237
/zg Cd/day
3.1
3.1
7.7
6.6
1.7
0.3
0.8
1.0
0.5
0.9
0.7
1.6
28.0
g/day
584
—
203
43
252
166
—
—
284
107
110
600
2,349
//g Cd/day
3.3
—
4.7
2.1
10.2
1.0
—
—
0.8
1.6
1.1
1.8
26.6
aFrom FDA Compliance Program Evaluation 1974 Total Diet Studies.
bAdjusted on a caloric basis from the FDA 1974 Total Diet Studies to represent the
normal diet which compares with the adult lacto-ovo-vegetarian diet.
°Loma Linda lacto-ovo-vegetarian diet. Based on response of 183 southern Californians
in a food frequency questionnaire bytheDepartmentof Biostatisticsand Epidemiology,
Loma Linda University School of Health, 1978. Leafy vegetables class includes root
vegetable and garden fruit classes from normal diet.
measured in Sweden, compared with a value of 10//g/day measured by the total diet
collection method. In Germany 31 //g/day has been measured, compared with 48
//g/day measured by the market-basket method. In Japan, where cadmium levels in
food are higher, the fecal analysis method has resulted in several estimates ranging
from 24 //g/day to 84 //g/day.
"It has generally been concluded that ingestion of 200 to 350 mg Cd/day
over a 50-year exposure period is a reasonable estimate for individuals
(excluding smokers and occupationally exposed) within the population to
reach the critical renal concentration (200 mg Cd/g wet weight in the renal
cortex) associated with the initiation of proteinuria. This ingestion limit
assumes background exposure levels of air and no exposure from smoking. If
these exposures are increased, then the suggested ingestion limit must be
correspondingly reduced. Smoking one pack of cigarettes/ day will reduce the
limit by about 25 //g/day. Again these exposures are assumed to occur over a
50-year exposure period and, in the case of cigarettes, since many smokers start
as teenagers, this addition would be relevant for much (30 to 35 years) of the
50-year exposure period. Therefore, smokers must be considered as being at
increased risk." (Ryan el al 1982).
26
-------
Thus, present levels of total dietary intake of cadmium for most people appear to
be fairly safe. However, in view of human variability in sensitivity and the variability
in food supply, these levels probably should not be allowed to rise greatly.
It is of interest to note that increased consumption by individuals of those leafy and
root vegetable crops highest in cadmium, and of organ meats as well, would increase
the dietary iron intake. Since iron-sufficient humans absorb only about 2.3% of
dietary cadmium, compared to an average absorption of 4.8% (in the generally
iron-deficient American population), the increased iron intake would tend to correct
for the increased cadmium intake (Chang 1980).
The potential increase in cadmium levels in human food due to land treatment or
irrigation is still an unsettled question. Almost all the relevant research on the subject
has been done with the land application of wastewater sludge, but, in spite of the high
cadmium-application rates associated with this practice, insufficient time has elapsed
to allow many firm conclusions to be drawn. It is clear, however, that increased
cadmium in the soil results in increased cadmium in the plants grown in that soil, the
degree of increase being a function of cadmium amendment, plant species and
cultivar, soil pH, organic matter, and time since application (Ryan el al. 1982). The
degree of risk to man, of course, is dependent upon the amount of the food supply
affected and the diet selection of the individual.
The most significant research need in the areas of trace elements probably
continues to be the development of an understanding of the factors controlling the
uptake of trace elements by plant crops at land treatment sites, and their entry into
the human food supply.
27
-------
NITRATES
Nitrogenous wastes are important constituents of municipal wastewaters,
consisting of (1) proteins and other nitrogenous organics from feces, food wastes,
etc., (2) urea from urine, and (3) their breakdown products. Raw domestic
wastewater has concentrations of about 8-35 mg/1 organic nitrogen, 12-50 mg/1
ammonium (plus ammonia), and, thus, 20-85 mg/1 total nitrogen, all expressed as N
(Metcalf and Eddy 1972). Nitrites and nitrates are normally present only in trace
amounts in fresh wastewater. Bacteria rapidly decompose most forms of organic
nitrogen to ammonium (or ammonia), in wastewater or soil. Under aerobic
conditions ammonium is oxidized by bacteria (Nitrosomonas) to nitrite, and the
nitrite rapidly oxidized by bacteria (Nilrobacter) to nitrate; the two-step process is
called "nitrification." Under anaerobic conditions,and in the presence of organic
matter, bacteria can use nitrate as a source of oxygen, and convert nitrate to
molecular nitrogen, which escapes to the atmosphere; this is called "denitrification."
Both aquatic and terrestrial plants can use ammonium and nitrate as a nitrogen
source.
Inorganic nitrogen is normally quite innocuous from a human health point of
view, although high ammonia levels can present an aesthetic problem. The major
health concern is that infants, less than about three months of age and consuming
large quantities of high-nitrate drinking water through prepared formula, have a high
risk of developing methemoglobinemia. The incompletely developed capacity to
secrete gastric acid in the infant allows the gastric pH to rise sufficiently to encourage
the growth of bacteria which reduce nitrate to nitrite in the upper gastrointestinal
tract. The nitrite is absorbed into the bloodstream, and oxidizes the ferrous iron in
hemoglobin to the ferric state, in which form it is incapable of carrying oxygen. Fetal
hemoglobin (Hb F), 50-89% of total hemoglobin at birth, is particularly susceptible
to this transformation. Methemoglobin is normally present in the erythrocytes of
adults, at a concentration of about 1% of total hemoglobin, being formed by
numerous agents, but kept to a low level by the methemoglobin reductase enzyme
system. This enzyme system is normally not completely developed in young infants.
At a methemoglobin concentration of about 5-10% of total hemoglobin the body's
oxygen deficit results in clinically-detectable cyanosis. As a result of epidemiological
and clinical studies (Shuval and Gruener 1977, Craun et al. 1981, Fraser and Chilvers
1981) a primary drinking water standard of 10 mg/1 of nitrate-nitrogen (i.e., nitrate
expressed as N) has been established (USEPA 1977) to prevent this condition from
developing.
Besides methemoglobinemia, there is also some concern about nitrates resulting in
the formation of carcinogenic N-nitroso compounds in the gut, but this phenomenon
probably involves higher concentrations than the 10 mg/1 water standard (Fraser et
al. 1980, Fraser and Chilvers 1981).
The relevance of land treatment, of course, centers on the possibility of highly
soluble nitrates reaching groundwater which may be used as a potable water supply.
Preapplication Treatment
Wastewater treatment, planned or otherwise, results in the rapid breakdown of
organic nitrogen to ammonium, and the oxidation of ammonium to nitrate, i.e.,
nitrification. Under anaerobic conditions, for example at the bottom of a wastewater
stabilization pond, some denitrification may occur, resulting in nitrogen loss from
the wastewater (USEPA 1983). In a pond some nitrogen may also be taken up by
algae (DiGiano and Su 1977), but this is not really lost from the wastewater.
28
-------
During treatment, organic carbon is constantly lost as carbon dioxide, so that,
while the C:N ratio in raw wastewater is about 5:1, the C:N ratio in secondary effluent
may be about 1:2. Since denitrifying bacteria require a source of organic carbon to
supply their energy needs, the level of organic carbon in wastewater is critical to the
removal of the highly mobile nitrate ion by denitrification in the anaerobic layers of
the soil. Where glucose is used as the carbon source the C:N ratio required by the
denitrification reaction is about 3.2:1. Thus, if groundwater is to be protected from
nitrate and wastewater application rates are going to exceed the rate of nitrogen
uptake by plants (if any), primary treatment alone, i.e., sedimentation, has been
recommended as preapplication treatment (Pound el al. 1978).
The influence of the C:N ratio on the degree of denitrification in the soil has been
demonstrated in numerous field and laboratory studies. For example, in studies on
the application of primary and secondary effluent to soil columns with and without
vegetation, Lance el al. (1980) observed the following nitrogen removals:
Vegetated Column Nonvegetated Column
Primary Effluent:
Secondary Effluent:
81.8% 45.6%
48.1% 28.5%
Increased nitrogen removal occurred with higher C:N ratio (i.e., primary effluent)
and the presence of vegetation.
Groundwater
The threat of nitrates to groundwater is less at slow-rate than rapid-infiltration
sites because of lower application rates and the presence of plant uptake. (Overland
flow systems should not affect groundwater.) Nitrogen concentrations in applied
wastewater and groundwater at several operating slow-rate systems are summarized
in Table 13; nitrogen is reported as total nitrogen, but in wastewater consists mostly
Table 13. Wastewater and Groundwater Nitrogen at Slow-Rate Land
Treatment Sites (mg/l as N)
Site
Dickinson, ND
Hanover, NH
Lubbock, TX
Roswell, NM
San Angelo, TX
Braunschweig,
Germany
Wroclaw,
Poland
Melbourne,
Australia
Santiago,
Chile
Total
Applied
Wastewater
12.1
27-28
12-15
36
29.8
49.2
42-45
51.3
35-38
Nitrogen Concentration
Affected
Groundwater
3.0
7.3
50
5-7
16.7
32.8
6-31
6.6
12-15
Control
Groundwater Reference
1.1 Benham-Blair et al.
1979a
— Jenkins & Palazzo
1981
8.5 Hineslyera/.
1978
3 Koerner & Haws
1979a
22.0 Hossner et al.
1978
1 1.6 Tietjen et al.
1978
— Cebula & Kutera
1978
— McPherson 1978
2 7 Schalscha et al.
1979
-------
of organic N and ammonium, and in groundwater mostly of nitrate. Of the U. S. sites,
only that at Lubbock appears to be overloaded, resulting in a groundwater
concentration of 50 mg/1 N.
Nitrogen concentrations at several operating rapid-infiltration systems are
summarized in Table 14. It is evident that many rapid infiltration systems result in
groundwater nitrate concentrations above the 10 mg/1 drinking water standard.
Many of these systems could improve their performance by decreasing the loading
rate (thus, allowing more time for within-soil treatment), increasing the C:N ratio
(particularly by using primary effluent instead of secondary), and/or optimizing the
flooding-drying sequence. Leach and Enfield (1983), for example, have found an
operating sequence of one-day flooding and one-day drying to be the optimum
regime for total nitrogen removal in their experimental system. If it is not feasible to
prevent the groundwater beneath a rapid-infiltration site from rising above the
nitrate standard, it may be possible to ensure that the water is not used as a drinking
water supply. Rapid infiltration basins could be located where groundwater flows
into surface waters, or the percolate could be recovered with wells or underdrains for
surface discharge or reuse in crop irrigation (Reed 1979).
It should be kept in mind that land treatment sites are not the only source of nitrate
in groundwater. Many groundwaters are naturally high in nitrates, e.g., that in the
vicinity of San Angelo, Texas (Table 13), and in urban areas on-site absorption fields
and lawn fertilizers have been shown to be sources of nitrates in groundwater (Porter
1980).
Conclusions and Research Needs
Land treatment systems, particularly rapid-infiltration, threaten to raise the
nitrate concentration in their underlying groundwater above the drinking water
standard of 10 mg/1 as N. This can be prevented, however, by proper siting and
management practice, e.g., using high C:N ratio wastewater, matching loading rate
to crop uptake (for slow-rate systems), and optimizing the flooding-drying regime.
These management practices and the agronomic factors controlling the entrance of
nitrates into groundwater are important research needs.
30
-------
Table 14. Wastewater and Groundwater Nitrogen at Rapid-Infiltration Land Treatment Sites (mg/l as N)
Total Nitrogen Concentration
Site
Boulder, CO
Brookings, SD
Calumet, Ml
Fort Devens, MA
Hollister, CA
Lake George, NY
Milton, Wl
Phoenix, AZ
Vineland, NJ
Loading
Rate (m/yr)
48.8
12.2
17.1
30.5
15.2
58
244
61
—
Applied
Wastewater
16.5
10.9
24.4
50
40.2
12
26.3
27.4
34-41
Affected
Groundwater
9-16
6.2
7.1
19.6
1.7-7.8
8
15.2
9.6
17-29
Control
Groundwater
—
—
—
—
0-9.9
—
6.9
—
2.9
Reference
Smith efa/. 1979
USEPA 1981
USEPA 1981
USEPA 1981
Pound et a/. 1978
Aulenbach 1979
Benham-Blair et a/.
1979b
Bouwer et a/. 1 980
Koerner & Haws
1979b
-------
SODIUM
Sodium may enter the groundwater beneath land treatment sites, just as nitrogen
does. For example, at the the Wroclaw, Poland, sewage farm, an untreated
waste water sodium content of 83-115 mg/1 was reflected in a groundwater content of
90-144 mg/l(Cebulaand Kutera 1978). This level may be compared with the sodium
content of U. S. drinking waters, 58% of which have 0-20 mg/1, and only 14% of
which have over 100 mg/l( White el al. 1967). A primary drinking water standard for
sodium has not been established (USEPA 1977).
The health significance of these raised levels in groundwater, and thus potentially
in drinking water, is unclear, particularly since drinking water sodium is a small
portion of total dietary intake (2-8 g/day). That a decreased intake of sodium can
reduce blood pressure in subjects with mild hypertension has been shown by several
clinical studies, e.g., Beard el al. (1982), but the overall significance of restricted
sodium consumption to preventing hypertension in the general public has been
recently challenged (Kolata 1982, Boffey 1982, Puska el al. 1983, Rouse el al. 1983).
Similarly, the evidence that high sodium consumption causes hypertension is
tenuous. However, increased groundwater sodium concentrations beneath land
treatment sites should be kept in mind as a possible future health concern.
32
-------
COMPARISON WITH CONVENTIONAL SYSTEMS
The comparison of the potential health effects of land treatment, caused by both
pathogens (Kowal 1982) and toxic substances, with those of conventional treatment
is necessarily highly subjective. Nevertheless, there are suggestions that land
treatment is at least equally protective of public health as conventional treatment.
A comparison of bacterial aerosol levels at conventional activated sludge plants
and a spray irrigation land treatment site has been performed by Clark et al. (1978),
to evaluate relative human exposure levels at these two types of facilities. They
concluded that airborne bacterial levels, as measured by fecal coliforms, appear to be
higher at the activated sludge plants than at the spray irrigation facility.
A broad comparison of health risks between activated sludge treatment and
slow-rate land treatment has been performed by Crites and Uiga (1979, Uiga et al.
1978). The comparison assumed:(l) A flow of 3 M gal/day of domestic wastewater.
(2) Activated sludge treatment is followed by disinfection and surface water
discharge. (3) Land treatment is preceded by aerated lagoon preapplication
treatment and storage, and followed by percolate water recovery using underdrains
and surface water discharge, with no disinfection. They arrived at the following
conclusions comparing the two systems:
I. If maintained and operated properly, both conventional and land treatment
systems provide a large measure of safety for public health. Slow-rate land
treatment offers greater protection against parasites, viruses, trace organics,
halogenated organics, trace elements, and nitrate.
2. Since adequate removal of parasite eggs and cysts require such measures as
filtration or long detention times in ponds or storage lagoons, land treatment
offers greater protection from health risks.
3. Land treatment systems, especially those with ponds and storage lagoons,
remove viruses to a higher degree than do conventional treatment and
disinfection systems.
4. Land treatment systems are less susceptible to failure or upsets than
conventional systems, especially for small systems.
In an overview of existing land treatment systems, Iskandar (1978) concluded that
"the potential health hazards from land treatment are no greater and probably less
than those from conventional treatment Although land treatment, like all known
waste treatment systems, has potential health hazards associated with it, these risks
can be kept to a minimum." The risks associated with alternative systems of
wastewater treatment and disposal, conventional as well as land treatment, should be
defined and compared. He considered it to be rather odd that land treatment instills
so much more public fear than conventional treatment.
33
-------
GENERAL CONCLUSIONS
This section contains the general conclusions and research needs from both this
report on toxic substances, and an earlier companion report on pathogens, "Health
Effects of Land Treatment:Microbiological" (Kowal 1982).
Types and Levels in Wastewater
The types and levels in wastewater of most pathogens are fairly well understood,
with the exception of viruses. Since only a small fraction of the total viruses in
wastewater and other environmental samples may actually be detected, the
development of methods to recover and detect viruses continues to be a research
need. The occurrence of viruses in an environmental setting should probably be
based on viral tests rather than bacterial indicators since failures in this indicator
system have been reported.
The tremendous number of organic compounds possibly present in wastewater,
together with their myriad health effects and poorly understood behavior in the
environment, represent a considerable potential for adverse health effects. Most of
these can probably be prevented by simple design and monitoring measures; this, of
course, would not be true in the case of discharges containing high levels of particular
chemicals.
It seems reasonable to conclude that cadmium is the only trace element likely to be
of health concern to humans as a result of land treatment of wastewater, with the
exposure being through food plants or organ meats.
Preapplicafion Treatment
The level of preapplication treatment required for the protection of public health
may be as little as properly-designed sedimentation at land treatment sites with
limited public access, where crops are protected by appropriate crop choice and
waiting periods, and groundwater is protected by appropriate hydrological studies
and application rate selection. Where protection of groundwater cannot be assured,
wastewater stabilization ponds should be considered for virus removal, but further
investigations into the survival of viruses, in these ponds is an important research
need, as is that of protozoan cysts. Because of potential contamination of crops and
infection of animals, slow-rate and overland-flow systems should have high removal
rate of helminth eggs. These relatively simple pretreatment requirements would be
appropriate for many land treatment systems in the U. S., e.g., for many slow-rate
sites where crops for animal feed are grown.
Preapplication treatment by storage lagoons may remove considerable quantities
of organics, but cannot be relied upon to efficiently remove all toxic organics,
particularly since most pretreatment design questions center on inactivation of
pathogens.
The recommendations made in the paragraphs below assume a minimum level of
preapplication treatment, i.e., properly-designed sedimentation. In situations with
greater public access (e.g., renovated water reuse on golf courses), shorter waiting
periods before grazing or harvest of crops (e.g., agriculture in arid areas), or threat of
groundwater contamination (e.g., shallow water table used as a drinking water
source), more extensive preapplication treatment may be required. This treatment
may consist of wastewater stabilization ponds, conventional treatment unit
processes, or even disinfection. The exact degree of pretreatment required for these
34
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situations is site-specific, and recommendations should be determined separately for
each system (Lance and Gerba 1978).
Aerosols
Because of the potential exposure to aerosolized viruses at land treatment sites, it
would be prudent to limit public access to 100-200 m from a spray source, unless the
effluent has been disinfected. At this distance bacteria are also unlikely to pose
significant risk. Human exposure to pathogenic protozoa or helminth eggs through
aerosols is extremely unlikely.
Suppression of aerosol formation by the use of downward-directed, low-pressure
nozzles, ridge-and-furrow irrigation, or drip irrigation is recommended where these
application techniques are feasible.
Although removal rates of organics from wastewater by aerosolization and
volatilization are high, exposure through this route is unlikely to present any
significant health effect.
Surface Soil and Plants
The survival times of pathogens on soil and plants are summarized in Table 15
(after Feachem el al. 1978). Since pathogens survive for a much longer time on soil
than plants, the recommended waiting periods before harvest are based upon
probable contamination with soil.
Aerial crops with little chance for contact with soil should not be harvested for
human consumption for at least one month after the last wastewater application;
subsurface and low-growing crops for human consumption should not be grown at a
land treatment site for at least six months after last application. These waiting
periods need not apply to the growth of crops for animal feed, however.
An important research need is the effect of drying of the soil between wastewater
applications on the survival of surface-soil viruses.
The levels of toxic organics likely to be present in soils at land treatment sites will
probably result in extremely low levels in above-ground portions of plants, but levels
in roots, tubers, and bulbs may present a health hazard.
The potential increase in cadmium levels in human food due to land treatment or
irrigation is still an unsettled question. It is clear, however, that increased cadmium in
the soil results in increased cadmium in the plants grown in that soil, the degree of
increase being a function of cadmium amendment, plant species and cultivar, soil
pH, organic matter, and time since application (Ryan et al. 1982). The degree of risk
to man, of course, is dependent upon the amount of the food supply affected and the
diet selection of the individual. Present levels of total dietary intake of cadmium for
most people appear to be fairly safe. However, in view of human variability in
sensitivity and the variability in food supply, these levels probably should not be
allowed to rise greatly.
Table 15. Survival Times of Pathogens on Soil and Plants
Soil
Pathogen
Bacteria
Viruses
Protozoa
Helminths
Absolute
Maximum
1 year
6 months
10 days
7 years
Common
Maximum
2 months
3 months
2 days
2 years
Plants
Absolute
Maximum
6 months
2 months
5 days
5 months
Common
Maximum
1 month
1 month
2 days
1 month
35
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The most significant research need in the area of trace elements probably continues
to be the development of an understanding of the factors controlling the uptake of
trace elements by plant crops at land treatment sites, and their entry into the human
food supply.
Movement in Soil and Groundwater
Properly designed slow-rate land treatment systems pose little threat of bacterial
or viral contamination of groundwater. Considerable threat of bacterial contamina-
tion exists, however, at rapid-infiltration sites where the water table is shallow,
particularly if the soil is porous. The survival of bacteria in groundwater, once they
get there, is poorly understood, and is an important research need.
Likewise, considerable potential for viral contamination of groundwater exists at
rapid-infiltration sites, and appropriate preapplication treatment or management
techniques should be instituted, e.g., intermittent application of wastewater. Until
then, groundwater drawn for use as potable water supplies should be disinfected. The
factors controlling the migration of viruses in soils, and the survival of viruses in
groundwater, are poorly understood, and are significant research needs.
Human exposure to pathogenic protozoa or helminths through groundwater is
extremely unlikely.
Toxic organics can enter the groundwater, particularly at rapid infiltration sites,
and the application and soil factors controlling this transport, together with the
factors governing their movement and decomposition within groundwater, are
significant research needs.
Groundwater is unlikely to represent a significant trace element threat except at
rapid infiltration sites.
Land treatment systems, particularly rapid-infiltration, threaten to raise the
nitrate concentration in their underlying grounrfwater above the drinking water
standard of 10 mg/1 as N. This can be prevented, however, by proper siting and
management practice, e.g., using high C:N ratio wastewater, matching loading rate
to crop uptake (for slow-rate systems), and optimizing the flooding-drying regime.
These management practices and the agronomic factors controlling the entrance of
nitrates into groundwater are important research needs.
Increased groundwater sodium concentrations beneath land treatment sites
should be kept in mind as a possible future health concern.
Animals
There appears to be little danger of bacterial, viral, or protozoan disease to animals
grazing at land treatment sites if grazing does not resume until four weeks after last
application. However, the role of animals in transmitting human viral diseases at
land application sites is poorly known, and is a research need. Removal of helminth
eggs during preapplication treatment should eliminate the potential of disease from
those long-lived parasites. The feeding of land-treatment-site-grown plants to
animals is unlikely to pose a health problem, but grazing animals may accumulate
significant levels of toxic organics. The issue of accumulation of organics from the
soil by plants and animals (particularly into milk) is poorly understood, and more
research is required.
Infective Dose, Risk of Infection, Epidemiology
Because of the possibility of picking up an infection, it would be wise for humans to
maintain a minimum amount of contact with an active land treatment site. The
comparison of the respiratory infective dose of enteric viruses with the oral infective
dose is a significant research need.
36
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Epidemiological studies to date suggest little effect of land treatment on disease
incidence. However, well planned and implemented prospective studies have not
been completed.
Perhaps most importantly, a comparison of land treatment systems with
conventional treatment systems suggests that the former is at least equally protective
of public health as the latter.
37
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