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Martin Maltempo
Daniel Chiras
John Morris
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Risk Assessment of
Wastewater Disinfection
BY
David Hubly
Willard Chapped
John Lanning
Martin Maltempo
Daniel Chiras
John Morris
UNIVERSITY OF COLORADO AT DENVER
DENVER, COLORADO 80202
TECHNOMICl
^PUBLISHING CO., INCy
LANCASTER-BASE!,
TECHNOMIC Publishing Company, Inc.
851 New Holland Avenue, Box 3535, Lancaster, Pennsylvania 17604, USA
TECHNOMIC Publishing AG
Elisabethenstrasse 15, CH-4051 Basel, Switzerland
ISBN 87762-517-4
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DISCLAIMER
The information in this document has been funded wholly or in
part by the United States Environmental Protection Agency under
assistance agreement number R-806586 to the University of
Colorado at Denver. It has been subject to the Agency's peer and
administrative review, and it has been approved for publication
as an EPA document. Mention of trade names or commercial products
does not constitute endorsement or recommendation for use.
11
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FOREWORD
The U.S. Environmental Protection Agency is charged by
Congress with protecting the Nation's land, air, and water
systems. Under a mandate of national environmental laws, the
agency strives to formulate and implement actions leading to
a compatible balance between human activities and the ability
of natural systems to support and nurture life. The Clean
Water Act, the Safe Drinking Water Act, and the Toxics Sub-
stances Control Act are three of the major congressional laws
that provide the framework for restoring and maintaining the
integrity of our Nation's water, for preserving and enhancing
the water we drink, and for protecting the environment from
toxic substances. These laws direct the EPA to perform
research to define our environmental problems, measure the
impacts, and search for solutions.
The Water Engineering Research Laboratory is that com-
ponent of EPA's Research and Development program concerned
with preventing, treating, and managing municipal and in-
dustrial wastewater discharges; establishing practices to
control and remove contaminants from drinking water and to
prevent its deterioration during storage and distribution;
and assessing the nature and controllability of releases of
toxic substances to the air, water, and land from manufactur-
ing processes and subsequent product uses. This publication
is one of the products of that research and provides a vital
communication link between the researcher and the user com-
muni ty.
Cost-effectiveness and environmental assessment are the
two criteria most commonly used to select among alternative
problem solutions; however, when substantive risk is created
by alternative solutions, risk assessment becomes a third
decision criterion. The research presented in this publica-
tion illustrates the application of risk assessment methods
to wastewater disinfection, a problem where all alternatives
create human and ecological risks.
Francis T. Mayo, Director
Water Engineering Research Laboratory
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ABSTRACT
A risk assessment data base is presented for several
wastewater disinfection alternatives, including chlorination,
ozonation, chlorination/dechlorination, and ultraviolet
radiation. The data base covers hazards and consequences related
to onsite use and transportation of the disinfectants and
ultimate disposal of disinfected effluents. A major segment of
the data base deals with the effects of chlorination products in
aquatic ecosystems. Energy consumption and cost analyses are also
presented for chlorination and ozonation. Example risk
calculations are presented for two hypothetical wastewater
treatment plants. The usefulness of the data base for riok
assessment is also discussed.
This report is submitted in fulfillment of Contract No.
R-806586-1 by the University of Colorado at Denver under the
sponsorship of the U.S. Environmental Protection Agency. This
report covers the period October 1979 to January 1984, and work
was completed as of January 1985.
IV
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CONTENTS
Foreword iii
Abstract iv
Figures. .• vi
Tables vii
Acknowledgments x
1. Introduction 1
2. Conclusion 3
3. Risk Assessment Methods 7
4. Chlorine
Hazard Identification 11
Severity and Frequency of Identified Hazards. . 18
Chlorination/Dechlorination 56
5. Hazard Identification for Ozone 60
6. Hazard Identification for Ultraviolet Radiation . 69
7. No Disinfection
Hazard Identification 72
Severity and Frequency of Hazards 76
No Disinfection Risk Model 80
8. Cost and Energy Considerations
Energy 87
Costs 94
9- Risk Model
Risk Model Data Base 105
Risk Model Examples 106
References 118
Appendices
A. U. S. Chlorine Producers and Packagers 151
B. Summary of Reported Chlorine Effects on
Freshwater Organisms 153
C. Summary of Toxic Effects of Chlorine to
Marine Aquatic Life 165
D. Summary of Chlorine Reaction Prod'uct Effects on
Freshwater and Marine Organisms 170
E. Summary of Toxic Effects of Ozone on
Aquatic Organisms 173
v
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FIGURES
Number Page
1 Location Map of U.S. Chlorine Producers 31
2 Responses of Selected Freshwater Organisms
to Total Chlorine Residual 35
3 Summary of Residual Chlorine Effects
on Freshwater Organisms 36
4 Comparison of Salmonella Dose-Response Models. . 85
VI
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TABLES
Number Page
1 1979 NSC Accident Data 21
2 1979 WPCF Accident Data 22
3 Summary of 1979 AWWA Accident Data 22
<4 Summary of 1979 AWWA Accident Data
by Category 23
5 1972 Colorado Accidents by Category 24
6 Compilation of DOT Accident Data 27
7 Percent Distribution of Chlorine
by Transportation Mode and by
Shipment Weight for 1972 28
8 Chlorine Shipments by Transportation Mode
andContainer 28
9 Accident Rates per metric ton-km 29
10 Chloroform Uptake from Fluids 42
11 Chloroform Uptake from Environmental Sources ... 42
12 Chloroform Uptake from Air, Water, and Food. ... 43
13 Carcinogenic Assessment of Chlorophenols in
Laboratory Species 50
14 Effects of 5-Chlorouraci1 on Spotted Sea
Trout Eggs and Larvae 54
15 Effluent Levels and Effects Ranges for Selected
Wastewater Chlorination Byproducts 55
16 Influent Concentration Ranges for Pathogenic and
Indicator Organisms 80
vii
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17 Microorganism Reductions by Conventional
Treatment Processes 81
18 Secondary Effluent Ranges for Pathogenic and
Indicator Organisms 81
19 Energy Utilization for European and Canadian
Plants Treating Municipal Water Supplies and
One U.S. Plant Treating Municipal Wastewater . 90
20 Energy Utilization for Oxygen Fed Wastewater
Ozonation Plant 91
21 Energy Utilization per Unit of Ozonated
Wastewater 92
22 Example Comparison of Energy Requirements for
Alternative Disinfectants 93
23 Construction Cost Indices 94
24 Chlorination Capital Costs 95
25 Annual Chlorination Capital Costs 95
26 Unit Chlorination Capital Costs 96
27 Chlorination Disinfectant Costs 97
28 Chlorination Labor Costs 97
29 Chlorination Power Costs 98
30 Chlorination O & M Costs Summary 98
31 Utilization Effects on Chlorination 0 & M Costs. . 99
32 Chlorination Costs Summary 99
33 Ozone Production Required to Produce an
Effective Dose of 5 mg/1 100
34 Ozonation Capital Costs 101
35 Annual Ozonation Capital Costs 101
36 Ozonation Unit Capital Costs 102
37 Ozonation Power Costs 102
38 Ozonation Labor Costs 103
Vlll
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39 Ozonation 0 & M Costs Summary 103
40 Ozonation Costs Summary 104
41 Estimate of Historical Data Quality 106
42 Effects of Example A Facility
on Aquatic Organisms Ill
43 Risks Summary - Example A 113
44 Effects of Example B Facility
on Aquatic Organisms 116
45 Risks Summary - Example B 117
IX
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ACKNOWLEDGMENTS
The authors wish to recognize the significant contributions
of several people not listed as authors. Terry Tedeschi was a
great help in initially organizing the project, and subsequently
she volunteered a su-bstantial amount of time for editing the
final draft of this report. During the data collection phase,
three students, David Shugarts, Robert Williams, and Barbara
Taylor, provided parts of the literature review. Larry Gratt
provided direction in selecting and using risk assessment
methods, and Jeffrey Feerer helped develop the example risk
calculations, finally, durinq the initial import writing phase,
Betty Lepthien, entered most «t the text into the word processor.
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SECTION 1
INTRODUCTION
Risks in today's world have been the center of growing
attention. Increased risk awareness in modern technological
societies is an outgrowth of technologic development and economic
achievement. Increased risk awareness may also be the result of
risk overload, a condition in which the maximum tolerance for
risk is being taxed.
Regardless of the reasons for an increased risk awareness,
technologies with inherent risks (air, water and soil pollution,
for example) cannot continue to develop without considering the
net impact of those inherent risks on humans, other living
organisms, and the environment. Also, cost-effective public
policy cannot be formulated without adequate assessment of risk.
Finally, because the world will never be free of risk, public
policy should be formulated with some consideration for its
management.
Effective risk management is based on a qualitative and
quantitative understanding of the risks associated with public
policy decisions. Risk assessment provides that understanding
and is the first step in risk management.
In 1979, faculty members at the University of Colorado at
Denver undertook the development of a wastewater disinfection
risk assessment because a controversy raged over whether
wastewater effluents should be disinfected and, if so, which
method was preferred. Since all of the alternatives included
inherent risks, the wastewater disinfection question was an
appropriate application of risk management.
A risk assessment of all alternatives seemed unnecessary and
inappropriate because so many disinfectants are unlikely to be
used. Thus, the assessment focused on chlorination and those
disinfection processes that appeared most likely to replace
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chlorination. The alternatives selected for the risk assessment
were as follows:
1. Chlorination,
2. Chlorination followed by dechlorination,
3. Ozonation,
4. Ultraviolet radiation, and
5. No disinfection
The choice of a disinfectant should be made at a regional or
local level because key variables such as the length of the
chlorine haul are site specific. Furthermore, since the water
quality management structure evolving in the United States
contains a strong emphasis on local and regional planning and
policy making, the use of this risk assessment data base in
evaluating disinfection alternatives is likely to occur at those
government levels. Thus the specific products expected from this
risk assessment are designed for local public policy-setting
applications. The two primary products are:
1. The collection and evaluation of a data base, and
2. The development of a method for using that data base
in a wastewater disinfection risk assessment.
A preliminary analysis of the risks associated with
wastewater disinfection revealed the broad mix of expertise
required to investigate such risks. Thus the study team was
designed to be interdisciplinary. Most of the work reported here
was developed by a physicist, a chemist, a biologist, an
economist, and a civil engineer.
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SECTION 2
CONCLUSION
A risk analysis is a detailed examination performed to
understand the nature of unwanted, negative consequences to human
life, health, property, or the environment. The purpose of this
analytical process is:|to provide information regarding the nature
and frequency of these negative consequences. The methods used
depend on the objectives, the data available, and the resources.
The analysis can range from a back-of-the-envelope calculation to
a fault-tree analysis of each element in the system(s). Each risk
calculation has an associated confidence interval that may or may
not be explicit in the calculations.
In most if not all such analyses, resource constraints and
data gaps make quantitative estimates of some risks impossible.
Such was the case with the study reported here. Nevertheless,
some useful conclusions have emerged.
1. Though wastewater treatment plants have a poor overall
safety record, with an accident rate similar to that of
metal mining, exposure to toxic substances (primarily
chlorine) accounts for only 4% of these accidents.
Note, however, that data on the relative severity (e.g.
lost workdays) of these exposures are not available.
Thus though the rate is similar to that for insect
bites, the risk (rate x severity) may not be similar.
In addition, the possibility of a low-probability,
high-consequence event (i.e., massive exposure) cannot be
discounted but is impossible to quantify without a more
detailed analysis (i.e. , a fault tree).
2. Most chlorine is shipped by rail (85%). The bulk of the
rest is shipped by tank truck (9.9%) or by common
carrier as cylinders (114 kg or 0.9 metric ton). Though
railroads generally have a much better safety record
than truck shipments (particularly in 114-kg
cylinders), the Youngstown, Florida, accident in
February 1978 illustrates the possiblity of low-
frequency, high- consequence events for railroad
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shipment. In this case, there were 8 fatalities, 260
injuries, more than $1 million in property damage, and
a release of 45,400 kg of chlorine gas.
3. A review of the literature on aquatic toxicology of
chlorine indicates that for all byproducts (e.g.
chloroform) except total residual chlorine (TRC) the
observed effluent levels are below those known to be
acutely toxic. Thus only TRC will lead to an acute
response, and then only if the dilution in the
receiving stream is insufficient to lower the resulting
level sufficiently. These effects can range from
avoidance to death of aquatic organisms. Further
studies are needed to determine whether TRC and various
byproducts could lead to chronic effects in aquatic
organisms at the levels encountered.
4. The compounds found in effluents are well below the
acute toxicity levels for humans. In general, the
contribution to finished drinking water from wastewater
disinfection will be much smaller than from drinking
water treatment. Though chloroform and
trichloroethylene are carcinogens and (assuming no
threshold) will therefore contribute some additional
risk of cancer to humans, this contribution is
relatively small (less than one excess case of cancer
for every 50,000 persons exposed to 5 ug/L of
chloroform over a lifetime). This assumes no dilution
of the effluent or subsequent loss. Since drinking
water chlorination is the major cause of human exposure
to byproducts of chlorination, these risks are not
expected to be an important consideration for
wastewater disinfection.
5. Though it is possible to identify hazards associated
with the alternatives of chlorination-dechlorination,
ozonation, and UV disinfection, the lack of data made
quantitation impossible. In general, these methods are
less of a concern for the environment and the general
public than chlorination. However, accidental releases
of S02 (used in dechlorination) can pose hazards to
humans and terrestrial and aquatic organisms. Ozone
poses a risk to workers in the plant and to vegetation
in the vicinity of the facility. Though Europe has had
considerable experience with ozone disinfection of
drinking water, no data on human risks are readily
available. Although ozone is toxic to aquatic life, its
lack of stability in water makes that risk minimal.
Even less information is available on the risk of
ultraviolet disinfection. The primary hazards are from
human exposure to the radiation itself (burns),
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exposure to ozone produced by the radiation, and
exposure to high electrical voltages (a hazard also
present with ozonation).
6. The risks of not disinfecting wastewater were also
analyzed. There was no discussion in the scientific
literature on the effect on aquatic organisms, though
it might be assumed that increased nutrients could have
some detrimental effects. The human hazards, of course,
relate to exposure to pathogenic organisms. The most
common risk would be gastrointestinal illness, although
there is also some risk of exposure to life-threatening
or disabling organisms as well.
7. An analysis was also performed on the energy use and
cost of chlorination versus ozonation. The capital
costs are about twice as great for ozone as for
chlorine, and operating and maintenance costs are, in
the best case, 35% higher for ozone than chlorine.
However, wide variation exists in the operation and
maintenance costs for ozone, depending on the
efficiency of ozone generation and absorption and the
cost of energy. Note, however, that the cost of
disinfection is only a few per cent of the total cost
of wastewater treatment, resulting in a maximum
difference of 10% for the total cost. The energy use
for chlorine was found to be one tenth that for ozone
(and similar to that for UV disinfection), but only on-
site energy use was taken into account. Though the
original generation of chlorine is, in fact, energy-
intensive, the fact that chlorine use for disinfection
is such a small portion of the total chlorine use is
likely to lead to a disproportionately smaller share of
increased energy costs.
8. A risk assessment data base for the wastewater
disinfection alternatives of chlorination, ozonation,
ultraviolet radiation, chlorination/dechlorination, and
no disinfection has been collected and reported here.
Portions of the chlorination data base are tabulated in
the Appendices, and the sources of the data base are
listed in the References section of this report. The
data base is heavily skewed towards the chlorination
alternative. This imbalance creates an excessive amount
of attention on the hazards of chlorination and may
create the illusion that the other alternatives involve
less risk. In addition, the nature of the data base is
not well suited to quantitative risk assessment because
many of the data do not support the development of
dose-response relationships for many acute responses
and for essentially all chronic responses.
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9- The study's goal of developing a risk assessment tool
for use by local and regional policy makers in
evaluating the desirability of requiring wastewater
disinfection has not been realized because the
available data base is insufficient to support such
analyses. Fault tree analyses could possibly fulfill
that goal, but that project must be undertaken in
future research projects.
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SECTION 3
RISK ASSESSMENT METHODS
Risk assessments consist of several parts, but they do not
all contain the same parts. The elements in each risk assessment
are governed by the nature of the risk being assessed, the
quantity and quality of available data, and the level of risk
assessment effort that can be supported by the available
resources.
The several parts of a risk assessment are designed to
satisfy one or more of the following objectives:
1. Completely define the potential hazards associated with
the subject activity.
2. Define and quantify the relationship between exposure
and severity for each of the hazards defined in item 1
above.
3. Estimate the frequency and/or probability of exposure
for each of the identified hazards.
4. Synthesize the results of the above three tasks into a
risk model.
While the first objective realized must be item 1 above, the
other three objectives are not necessarily pursued or realized in
the order shown. For example, the type of data available is
usually discovered while pursuing items 2 and 3, and the results
of those tasks will determine the type and complexity of the risk
model chosen in item 4. However, the type of model chosen wil 1
provide the focus for the data searches in items 2 and 3. This
feedback relationship requires a risk assessment to use an
iterative method in order to achieve the highest level of clarity.
The hazard definition tasks are qualitative,and are
designed to answer such questions as, "What happens if humans (or
fish...or plants..etc.) are exposed to gaseous chlorine (or
dissolved chlorine...or chlorine reaction products...or
ozone...etc.)?". These questions can usually be answered through
laboratory or field studies and/or through extensive literature
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searches. There have been many studies designed to answer this
type of question so the literature search method can be very
fruitful. Sometimes it is necessary to identify intermediate
causes of hazards before this task can be completed. In this
assessment, hazards caused by process reaction products and the
identification of those hazards could not be accomplished until
the reaction products were identified. Therefore, some of the
hazard identification sections in this report also contain a
discussion of the expected process reaction products.
The quantitative elements of a risk assessment are produced
by the severity and frequency definition tasks. When dealing with
hazards caused by chemical agents, the severity definition
element usually involves the discovery or development of dose-
response models. Since chlorine, ozone, and the process reaction
products all fall into the "chemical agent" category, a major
part of this study involved the search for dose-response models.
The frequency definition task is dependent on the nature of
the available data. For example, transportation accident data are
often reported in terms of ton-miles, so the frequency definition
task must estimate the expected irureas^ or decrease in ton-miles
resulting from a change in policy or practice. This estimate can
then be transformed into expected accidents. Sometimes the
frequency definition task cannot be accomplished in a global or
generalized context. In those cases the frequency definition task
has been omitted from this study due to the narrowness of the
study scope.
The frequency definition task may also depend on the chosen
risk model. For example, if a risk model is chosen where the
hazard (or outcome) is a function of event A then the frequency
definition task must find means of predicting the occurrence of
event A. On the other hand, if the hazard is modeled as a
function of some measureable but not necessarily causative
quantity, such as man-years of labor, then the frequency
definition task is the same as in the paragraph above. The
dependence of the frequency definition task on the choice of risk
model should be clarified in the following discussion of risk
models.
Risk assessments have developed many ways of modeling risk;
however, all of the risk models attempt to do one thing: predict
the probability of realizing a hazard (or outcome) Y as a
function of an event A (or events A, B, C, etc....). Ideally the
model should be based only on cause and effect relationships,
i.e. hazard Y should be caused by event A. This model might be
shown mathematically as:
Y = f(A)
Such a simple relationship seldom occurs in natural systems.
8
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Usual ly the event A is also a function of an earlier event A^/
and that event in turn is a function of an even earlier event A2r
and so forth. This latter model would take the form:
Y = f (A),
Where:
A = f
A! = f (A2)
n-
The conversion of this latter cause and effect model to a
probability model leads to:
= «A/A1XA1/A2> ...... >
where is the probability that hazard Y will occur, and
, , ...., are the probabilites that
each of the causative events will occur given that the
previous event has occurred, and is the probability
that the initial event occurred.
This sequential event probability model is frequently used
in risk assessment because it usually is easier to define or
estimate the relationship between each pair of sequential events
than to define the relationship between the initial event
and the resultant hazard . This type of model requires that
all of the intermediate cause and effect relationships and the
severity and frequency elements of those relationships be defined
during the risk assessment.
The type of model becomes even more complex and costly if
hazard Y can be caused by more than one string of sequential
events. This more complex probability model for two strings of
sequential events would be:
+ ...... '
n 1 ...... n
where the term definitions are analagous to the equation above.
Further complexity is added when some of the sequential events
are also caused by more than one string of earlier events. Some
systems also contain synergistic and antagonistic relationships
between parallel event strings. Feedback relationships may also
exist.
These sequential event cause and effect models can be
plotted graphically. The direction of the effect and feedback
relationships are shown by lines and arrows, and the synergism
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and antagonism elements are shown with Boolean logic symbols.
The resulting plot resembles a tree, and this visual effect led
to the identification of this model as "fault tree analysis".
The high cost of fault tree analysis is usually justified
when existing data are not sufficient to construct a probability
relationship between the initial event(s) and the hazard Y. Fault
tree analysis permits the construction of such a probability
relationship by estimating probabilities for several simpler and
probably better known relationships. Analogies are often used to
estimate these simpler probability relationships.
However, because the funds available were not sufficient to
produce a fault tree analysis, this risk assessment was limited to
the simplest risk model described above. The funding constraint
also eliminated consideration of risks associated with production
of the disinfectants at sites other than the wastewater treatment
plant site and consideration of indirect risks such as the risks
associated with the production of power to operate the
disinfection processes. These constraints resulted in a different
product than had first been envisioned.
10
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SECTION 4
CHLORINE
HAZARD IDENTIFICATION
On-site Use Hazards
Although the use of chlorine to disinfect wastewater
effluents is an effective mechanism for the destruction of
pathogens to accepted levels, various risks or hazards can be
attributed to chlorine. In addition, there is a growing concern
and awareness about the presence of chlorinated chemicals and
their associated risks in wastewater. Chlorine is of obvious
concern because of its widespread use as a wastewater
disinfectant. The identification of risks associated with
chlorine as a wastewater disinfectant can be grouped according to
production, transportation and handling, and use. Secondary risks
associated with power consumption in the production phase and fuel
consumption in the transportation phase can also be identified.
The use of chlorine as a wastewater disinfectant has very little
effect on the production level of chlorine, so if chlorine were
totally eliminated as a wastewater disinfectant, the decrease in
risk associated with the overall production of chlorine would be
small. Secondary risks by their very nature have less
significance. So neglecting these production and secondary risks
will not have a major impact on the overall risk analysis
project.
The primary risks associated with the handling of chlorine
are:
1. Human exposure to liquid chlorine
2. Human exposure, both occupational and public, to
gaseous chlorine
3. Vegetation exposure to either liquid oi. gaseous
chlorine
Exposure to liquid chlorine is possible for the occupational
workforce and can result in severe skin or eye burns. However,
11
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the most common exposure is to gaseous chlorine, which is the
normal chlorine state at atmospheric pressure and normal
temperatures.
The most important human exposure routes to gaseous chlorine
are inhalation followed by eye and skin contact. The odor
threshold of chlorine is approximately 0.2 ppm (parts-per-million
by volume). Below 1 ppm there is little dose-response correlation
for workers chronically exposed to chlorine, while chronic
exposure to 5 ppm chlorine can bring about respiratory
complaints, nausea, increased susceptibility to tuberculosis, and
corrosion of the teeth (NAS, 1976). Acute exposure to chlorine
presents immediate and latent effects. Immediate effects begin
with throat and mucuous membrane irritation for an exposure of
one hour at 7 ppm. Higher concentrations lead to cough,
conjunctivitis, pulmonary edema, and death. A 100 ppm exposure is
lethal in only a few seconds, representing a significant risk to
humans for an accidental release of gaseous chlorine. Latent
effects are less pronounced and often difficult to diagnose but
may include broncospasm, especially in asthmatic people, and
difficult or painful breathing (NAS, 1976).
There is a wide range of sensitivities among vegetation
species when exposed to gaseous chlorine. Chlorine exposures
result in spotting of vegetation at low concentrations to decay
and death of the plant at higher concentrations. Threshold
concentrations for acute injury vary but typically begin in the
0.5 to 1 ppm range for one hour exposures. The pattern of plant
injury for exposure to chlorine is similar to that by ozone or
sulfur dioxide in many species (NAS, 1976 ).
Transportation Hazards
The primary risks associated with the transportation of
chlorine are:
1. Human exposure to liquid chlorine
2. Human exposure, both occupational and public, to gaseous
chlorine
3. Human injuries incurred without contact with chlorine
The quantification of risks associated with the
transportation and handling of chlorine is made difficult by the
variety of transportation modes available,and the wj.de range in
sizes of shipping containers. Nevertheless, government
regulations associated with transporting dangerous chemicals such
as chlorine dictate an adequate data base for the risk analysis
project.
12
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Hazards of Process Reaction Products
The reaction products of the chlorination process that are
hazardous are soluble organic and inorganic species. As a result
the hazards are almost exclusively realized in the aquatic
environment receiving the disinfected wastewater and in
subsequent uses of those receiving waters.
Expected Reaction Products—
An assessment of risk associated with chlorine reaction
products should identify and analyze all compounds formed or
those whose concentrations increase during the disinfection
process. However, the number of potential reaction products is
large, so a limited number of compounds formed during the
disinfection process were selected for risk assessment. Selection
was based on reported effluent concentrations, toxicity, and the
availability of published data.
Although several studies have identified a large number of
chlorinated hydrocarbons in chlorinated secondary wastewater
effluent (Glaze, 1975; Glaze, 1973; Jolley, 1975; Jolley, 1979;
Environmental Protection Agency, 1979), selection of a small
number of representative or model compounds for the risk analysis
project has been difficult because of limitations within the
published data bases.
Studies on chlorination reaction products are typically
based on laboratory conditions which in turn often employ high
chlorine doses and/or long contact times , neither of which are
realistic when compared to normal treatment plant operating
conditions. The daily, weekly, and seasonal variations in plant
operating conditions add further complications in applying
laboratory data to field conditions. Additionally, most studies
used analytical techniques which were not effective in detecting
nonvolatile compounds. Only recently has there been much effort
placed on the difficult problem of separating and identifying
nonvolatile chlorinated hydrocarbons (Jolley, 1979). Although a
large number of chlorinated hydrocarbons have been separated in
chlorinated wastewater effluent, not all have been unambiguously
identified. A pilot study by the U.S. Environmental Protection
Agency is studying chlorinated effluents under actual field
conditions; however, the study is focusing on priority pollutants
and does not attempt to identify all chlorinated hydrocarbons
(Environmental Protection Agency, 1979a). Furthermore, toxicity
data are lacking for most chlorinated hydrocarbons thus far
identified in wastewater effluents.
13
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The compounds selected for risk analysis are:
1. chlorine residuals
(e.g. total residual chlorine (TRC)),
2. chlorinated aliphatic hydrocarbons
(e.g. chloroform, trichloroethylene,
and tetrachloroethylene), and
3. chlorinated aromatic hydrocarbons
(e.g. dichlorobenzene, chlorophenol,
and 5-chlorouracil).
Total residual chlorine was chosen because its concentration in
chlorinated wastewater effluents is orders of magnitude higher
than the chlorinated hydrocarbons (mg/L range versus ug/1 range),
and it is known to be toxic to aquatic life (Ward, 1978).
Compounds in the second category represent low molecular weight,
volatile chlorinated hydrocarbons. Although there is some
evidence to suggest chlorination actually lowers the chloroform
concentration in wastewater effluents (Environmental Protection
Agency, 1979a), chloroform was chosen because of i'Lsi relatively
high concentration in chlorinated wastewater effluents. Tri- and
tetrachloroethylene were chosen as representative volatile
chlorinated hydrocarbons for which there was a reasonable amount
of published toxicity information. Although other volatile
compounds could have been chosen, most have very limited
published toxicity information. The selection of chlorinated
nonvolatiles given within the third category was more difficult
due to very limited published information. Although the
nonvolatiles seem to be present at the low parts-per-bil lion
level, some are quite toxic and may represent significant risk to
both humans and aquatic life. The choice of chlorobenzenes and
chlorophenols was suggested by the availability of published
toxicity data, and 5-chlorouracil was chosen because other
investigators have selected this compound as a model for
compounds typically found in chlorinated effluents (Gehrs, 1974).
Hazards to Freshwater Fish and Invertebrates—
Chlorinated wastewater effluents containing a variety of
organochlorine compounds, total residual chlorine, and other
inorganic and organic molecules, may have acute (short-term)
toxic effects on aquatic organisms. Acute toxicity is generally
caused by high concentrations of a given chemical during short-
term exposures. The effects of these exposures are manifest
immediately and often subside when the exposure is eliminated. In
wastewater treatment plants acute exposures to high levels of
toxic chemicals would generally coincide with opening a new
plant, starting up after a shutdown, overchlorination during low
wastewater flow or periods of low receiving stream flow.
14
-------
Reactions to acute exposures of this nature could result in fish
kills and death of invertebrates in extreme cases or slight
physiological alterations in mild cases.
It is important to note, however, that acute toxicity may
not always be caused by high concentrations of chemicals in the
effluent. It is possible that highly toxic chemicals may exert
their effects at low levels over a short time span.
Acute toxic effects range in severity from slight,
reversible physiological perturbations to death. High levels of
toxic chemicals are generally associated with acute toxicity,
although the differential toxicity of chemicals precludes posing
this as an inviolable rule. It may be more appropriate to talk of
immediate effects rather than acute toxicity. Thus two types of
immediate effects can be identified:
1. those caused by low levels of a chemical or
chemicals, and
2. those caused by high levels.
Chronic toxicity includes effects which generally become
manifest after considerable time. Induction of tumors is a
classic example of a long-term or chronic effect. Concern for
chronic effects is generally centered on exposure to low levels
of pollutants over long periods. However, short-term exposures
may also have long-term effects, i.e., a single exposure may not
have an acute effect but may become manifest after a considerable
period of time after the exposure. Chronic toxicity is generally
discussed in reference to chemicals to which organisms are
exposed for long-time periods. The effects of such exposures are
not immediately apparent in most cases, but instead become
manifest years from the commencement of exposure. Short-term
exposures may also have a long-term, or delayed effect.
Therefore, it is best to refer to delayed effects from both long-
term and short-term exposure, rather than to use the term
"chronic toxicity".
Exposure to chemicals in the long- or short-term may have a
number of effects on aquatic organisms:
1. Alteration of normal physiological processes,
2. Induction of genetic mutations (mutagenesis),
3. Induction of cancer (carcinogenesis),
4. Induction of defects in offspring (teratogenesis),
5. Reproductive impairment of sexually mature individuals,
15
-------
6. Decreased survival of eggs, embryos, and other life
forms, and
7. Death.
These effects can drastically alter population size and may
result in serious upsets in the food web. For example, loss of
one species may result in subsequent losses in other species
which are dependent on the first as a food source. These indirect
effects are important in determining the overall impact of
wastewater chlorination on aquatic organisms. This risk
assessment focuses on the direct toxic effects of chemicals on
fish, invertebrates and some aquatic plant-life (notably algae).
It should be emphasized, however, that indirect effects (e.g.,
interruption of the food chain) may have serious consequences for
aquatic organisms.
An additional consideration in assessing the impact of
wastewater chlorination is the possibility that a chemical at a
given concentration may not affect adults but may be lethal to
eggs or finger lings. This age-related differential sensitivity is
an important factor to be analyzed in setting safe standards for
aquatic pollutants.
Behavioral changes, while somewhat ignored by researchers
and difficult to measure, could potentially occur in aquatic
organisms exposed to pollutants. Reproductive behavior and
migration, for instance, might be altered, thus affecting
population size as well as survival. Crowding behavior and
avoidance behavior (i.e., avoiding chlorinated water) have been
reported in the literature. Their significance, however, to the
overall stability of the aquatic ecosystem is not clear.
Hazards to Humans—
Humans consume surface water that, in some instances,
contains chlorinated organics and chlorine residuals. These
chemicals come from water chlorination at water and wastewater
treatment plants. Exposure to these chemical species may have
toxic effects. Immediate effects could range from slight
alterations of normal physiology to severe consequences such as
death. In reality, however, acute toxic effects are unlikely to
occur from consumption of water containing total chlorine
residuals and chlorinated organics because levels are generally
quite low. The concern for human health is generally focused on
long-term effects such as mutagenicity and carcinogenicity.
Concern for mutagenicity and carjinogenicity of water
chlorination by-products seers fairly well justified from the
scientific literature. The mutagenic potential of water
chlorination by-products deserves careful attention because of
the seriousness involved in the alteration of the human gene pool
16
-------
and because of the large population potentially at risk.
Teratogenic effects may also be manifest in human populations
exposed to water chlorination by-products.
Additional Complications—
Chemicals may interact in a synergistic manner causing
biological effects to be amplified. Synergism must be considered
at all times in risk assessments because of the potential
seriousness of the amplified effect. Analysis of studies on
individual chemicals may indicate that none has a significant
effect in low concentrations. However, exposure to several
chemicals together at low levels may give a biological response
which is greater than the sum of the effects when each chemical
is given alone. Synergism cannot be predicted and unless studies
are carried out with combinations of chemicals at low levels, so
conservatism is required in predicting the risk of wastewater
disinfection, especially with regard to mutagenesis,
teratogenesis, and carcinogenesis.
It is also possible that chemicals may antagonize one
another, i.e., the effects cancel out. The literature rarely
refers to this potential, but it is possible that chlorinated by-
products may have an antagonistic effect. In this case, the
response of two chemicals together is less than the sum of their
responses when administered alone.
Humans and aquatic organisms may be exposed to waterborne
chemicals from a variety of sources other than wastewater
disinfection. These sources of exposure should be considered
carefully when choosing a wastewater disinfection alternative.
In other words, the risk of a wastewater disinfection alternative
should be integrated into the total risk picture in such a way
that the total exposure to harmful chemicals can be made. This is
especially significant since synergism and antagonism may occur
with other environmental chemicals; that is, water chlorination
by-products may synergize or antagonize with chemicals which
enter an organism from air, food, medicine, etc.
An additional concern about the risks associated with
wastewater disinfection reaction products is bioaccumulation.
Bioaccumulation or biological magnification defines a phenomenon
that occurs in food chains. Certain chemicals (e.g., DDT,
chlorophenols, mercury) found in aqueous environments in low
concentrations tend to accumulate in organisms. As one ascends
the food chain, levels of these chemicals in the organisms
increase. At the top of the food chain, concentrations are the
highest and are often many thousands of times higher than in the
aqueous environment. Thus, seemingly low levels of a chemical in
water become hazardous since the organisms higher in the food
chain have accumulated the chemical. Large concentrations of a
chemical in the higher organisms may have a significant
biological effect.
17
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SEVERITY AND FREQUENCY OF IDENTIFIED HAZARDS
On-site Use Hazards
The goal for determining risk associated with the on-site
use of chlorine as a disinfectant is to establish an accident
frequency and the resulting severity rate for a variety of
wastewater plant sizes. After reviewing a variety of data bases
containing on-site accident information, the above goal was
virtually impossible to attain. Thus, the on-site accident
information for the use of all disinfectants has been listed as a
data deficiency for this risk analysis project. This deficiency
for on-site accident information has also been recognized by the
Water Pollution Control Federation.
The data bases reviewed for on-site accident information
will be discussed and shortcomings for the purpose of this risk
analysis project will be highlighted. Finally, the data bases
which allow an estimate of the risk associated with chlorine as a
disinfectant will be reviewed, and a procedure for estimating on-
site risk due to chlorine as a disinfectant will be developed.
United States Department of Labor—
Several departments, or bureaus, in the U.S. Department of
Labor compile accident statistics for wastewater treatment plant
workers. These include the Bureau of Labor Statistics, the Safety
Programs Office and the Office of Management Data Systems, both
of which are under the Occupational Safety and Health
Administration (OSHA).
The" annual accident statistics publications by the Bureau of
Labor Statistics are quite general, do not break down the workers
into a detailed four-digit Standard Industry Classification
(SIC), and do not contain the causes of accidents. In these
respects, these reports appear to be quite similar to those of
the National Safety Council.
The Office of Management Data Systems and Statistical
Analysis indicates four fatalities (three separate incidents)
which occurred under SIC 4952 (sewerage systems) for the period
7/72 to 4/80 that were investigated by OSHA. Although the causes
are not explicitly stated, standards from Title 29 of the Code of
Federal Regulations are cited , and the severity of the citation
is indicated. Further investigation may have revealed the causes
for the accidents; however, most wastewater treatment plants are
municipally owned and are not under OSHA requirements, thus
limiting the OHSA data base severely.
In cooperation with state and local governments, the U.S.
Department of Labor Bureau of Labor Statistics initiated the
Supplementary Data System (SDS) to enhance the mechanism for
collecting, coding, and analyzing statistical data concerning
18
-------
injuries and illnesses to workers. Although the SDS system can
provide detailed accident information on wastewater treatment
plant workers (or more precisely employees working on or for a
wastewater treatment plant), the SDS data have proven to be
unreliable for risk analysis purposes.
Although the exact reporting systems vary from state to
state for SDS, the data are very detailed and suitable for
determining the cause and severity of an injury for risk analysis
purposes. Primary data from a typical SDS include:
1. Date of accident, date of claim, and/or claim number
2. Occupation of injured worker within a given SIC
3. SIC
4. Source of injury
5. Nature or type of injury
6. Severity of injury or extent of disability
Information required for the risk analysis project which is
missing from the SDS include:
1. Treatment capacity of the wastewater treatment plant
2. Number of workers in SIC and plant or number in SIC
state-wide
3. Number of employee-hours in SIC at plant or number in
SIC statewide
Although SDS reports are open to the public, some states
code information to prevent tracking accidents to the employer or
individual treatment plant. When permissible, it may be possible
to obtain some of this missing information by tracking accidents
back to the specific treatment plant. In addition, some of the
missing information could be obtained from other state agencies.
Missing information for SDS data was not pursued since,
although the SDS is a good source of detailed information, it has
some problems when trying to obtain accurate and reliable
accident information for the risk analysis project. A brief
review of thes- -'roblems is given below:
1. There is apparently no federal requirement to belong to
the SDS. In the western part of the United States, North
Dakota, Kansas, Oklahoma, Louisiana, Texas, Nevada, and
Illinois do not belong to the SDS while South Dakota is
reportedly dropping out of the system.
19
-------
2. Standardized reporting guidelines by the Department of
Labor do not exist. Consequently, the type of
information gathered, and reported, varies from state to
state. Even though most states have computerized SDS
data, obtaining data printouts is often time consuming,
and the different coding systems and resulting different
reporting formats dictate an explanation booklet
accompany each state's data printout.
3. The SDS is typically coupled to the state workmen's
compensation fund. Therefore, not all accidents are
required to be reported. Accidents which result in less
than a set number of lost workdays need not be reported.
This "grace period" varies from state to state with a
range of zero to seven days. A three to four day grace
period is typical and determining the number of lost
workdays is open to some subjective interpretation by
employers. When trying to obtain reliable accident rates
regardless of severity, this grace period is an obvious
problem.
4. Data are open to bias from both the employer and state
agency. Information for the SDS is compiled from reports
submitted by employers concerning accidents of their
employees. Employers may underestimate the type and
severity of injury, may misrepresent the source of the
injury, or may provide incomplete or ambiguous
information (Colorado Division of Labor, 1978). State
personnel who collect and compile the accident
information may not be experts in classifying, or
coding the information provided. The term NEC (Not
Elsewhere Classified) is a category which is all too
often encountered. In both cases, the reliability of the
SDS data is open to question.
National Safety Council—
The National Safety Council (NSC) publishes Accident Facts,
which is an annual publication that includes accident rates by
SIC code. Unfortunately, wastewater treatment plant workers
require a four digit SIC code and Accident Facts compiles
information into two and three digit SIC codes. SIC grouping 495
is not one of the selected industrial groupings in Accident
Facts.
Another publication by NSC, Work Injury and IIIness Rates,
includes .1 breakdown of injury and illness incidence rates by SIC
code and industry. The 1980 edition is the first to include the
4952 SIC. The 1980 data covering the period 1977-1979 are
obtained from only three reporting units, which are the minimum
number of units required for publishing data. The usefulness of
this limited data base is questionable.
20
-------
Also contained in Work Injury and IIIness Rates (National
Safety Council, 1980) are data for SIC 4952 which the NSC
collected in cooperation with the American Public Works
Association. The data are collected from a substantially larger
base and are summarized in Table 1 along with the NSC comparison
data for transportation and public utilities as well as all
industries.
TABLE 1. 1979 NSC ACCIDENT DATA
SIC 4952
SIC 4(transpc
reporting
units
156
)rtation & 1290
lost work-
day cases
8.21
3.24
lost
workdays
115
63
public utilities)
All Industries 8377 2.67 61
Note - lost workday cases and lost workdays are calculated using
a base of 200,000 man-hours.
Water Pollution Control Federation—
The Water Pollution Control Federation (WPCF) compiles
statistics on accidents occurring to wastewater treatment plant
personnel. Results of the WPCF 1980 Safety Survey for accidents
occurring in 1979 were published in the December 1980 issue of
Deeds and Data (Water Pollution Control Federation, 1980). The
WPCF data are based on 1422 survey responses which represent a
substantial 7 to 10 percent of all wastewater works in the U.S.
and Canada. Not only does the report use a substantial data base,
but wastewater works are broken into collection and treatment
facilities. The survey also includes the number of employees and
man-hours for the reporting units. The WPCF data are summarized
in Table 2.
The 1980 WPCF survey, as in all previous surveys, does not
include information as to the cause or nature of the injury. The
1981 survey, which is to be compiled late in 1981, will include
questions on the cause and type of injury. WPCF has recognized
the need for more accident information on wastewater treatment
plant personnel (Hadeed, 1981), and this detailed accident
information would be most helpful in the risk analysis project.
21
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TABLE 2. 1979 WPCF ACCIDENT DATA
plant size
m^ per day
<3790
3790-9475
9475-37900
>37900
man-hr per
employee
1861
1985
1945
1958
injury
frequency
/106 man-hr
22.16
38.99
48.23
61.95
severity rate
lost workdays
/106 man-hr
252.9
210.9
436.9
749.3
fatalities
/106 man-hr
0
0
0
0
average
1952
52.48
566.1
Notes
1. Data are for treatment plants only and do not include data
for collections systems.
2. Plant size calculated assuming 379 liters per person per day.
American Water Works Association—
Although the American Water Works Association (AWWA) does
not collect accident data for wastewater treatment plants, they
do collect accident data for personnel at drinking water
treatment plants. The AWWA data include injury frequency and
severity rate for a parallel industry which uses chlorine as the
TABLE 3. SUMMARY OF 1979 AWWA ACCIDENT DATA
units
Man-hr per
employee
Injury
Frequency
/106 man-hr
Severity
lost workdays
/106 man-hr
Totals
2611
2021
35.19
1408
Notes
1. AWWA data include collection, treatment, and distribution
phases of the water treatment industry for U.S. and
Canada.
disinfectant. The use of AWWA data appears to be valid since both
water and wastewater treatment facilities -_tre largely
municipally owned, both industries use ^he same disinfectant and
disinfection equipment, both industries show similar ranges in
plant size, and both industries have about the same degree of
automation. Furthermore, the AWWA data include information which
is not available from the wastewater industry; namely, a
breakdown of accident types. AWWA data are summarized in Tables 3
22
-------
and 4 (American Water Works Association, 1979). The percentage
figures shown in Table 4 are based on the frequency of
occurrence, not on the severity measure, lost workdays.
TABLE 4. SUMMARY OF 1979 AWWA ACCIDENT DATA
BY CATEGORY
INJURY PERCENT
1. Sprain/strain in lifting, pulling, or pushing 24
2. Sprains/strains due to awkward position sudden 16
twist or slip
3. Struck against stationary or moving objects 11
4. Struck by falling or flying objects 11
5. Falls on same level to working surface 7
6. Caught in, under, or between object 7
7. Falls to different level from platform, 6
ladder, stairs, etc.
8. Contact with radiations, caustics, toxic, 4
and noxious substances
9. Animal or insect bites 3
10. Rubbed or abraided 1
11. Contact with temperature extremes 1
12. Contact with electric current 0
13. Miscellaneous 10
For the purpose of comparison, accident types for disabling
injuries from a state-wide industrial base for the State of
Colorado are summarized in Table 5 (Colorado Division of Labor,
1978).
Before differences or similarities in the data can be
examined, differences in the reporting systems must be
considered. The National Safety Council uses the record keeping
requirements found in the Occupational Safety and Health Act of
1970 (OSHA format) while the American Water Works Association and
the Water Pollution Control Federation use the American National
Standard Institute's (ANSI) method of recording and measuring
work injury experience, ANSI Standard Z16.1-1967 which was
reaffirmed in 1972 (ANSI Z16.1 format). Definitions for and
differences between the two systems are noted by the National
Safety Council (National Safety Council, 1980). The major
differences are:
1. Z16.1 uses 1,000,000 man-hours while OSHA uses 200,000
man-hours to calculate incidence rates.
23
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TABLE 5. 1977 COLORADO ACCIDENTS BY CATEGORY
Category Percent
1.
2.
3.
4.
5.
6.
7.
8.
9.
10.
11.
12.
Overexertion
Struck by
Struck against
Fall on same level
Bodily reaction
Caught in, under, or between
Fall from elevation
Rubbed or abraded
Motor vehicle accidents
Contact with radiation or caustics
Contact with temperature extremes
Contact with electrical current, public
transportation, NEC
25.0
19.9
13.0
9.7
6.2
7.5
7.5
2.4
2.9
3.0
2.2
0.6
2. Z16.1 measures ability to work whether scheduled to work
or not (calendar days), while OSHA measures only
scheduled workdays (five day work week) to calculate lost
workdays.
3. OSHA has no schedule of time charges for deaths and
permanent injury.
Since most wastewater treatment plants are municipally owned,
OSHA record keeping is not required and the Z16.1 format is
typically used.
Conclusions—
Data from the National Safety Council, the American Water
Works Association, and the Water Pollution Contol Federation are
useful for the purpose of risk analysis. The conclusions are
summarized below:
1. Personnel employed at wastewater treatment plants have a
poor overall safety record. The NSC SIC 4952 data compare
quite we 1 1 to that of the WPCF once the base man-hour
difference is considered. These data point to an accident
rate considerably above the average for all industries.
The accident rate, in terms of total lost workdays for
SIC 4952, is comparable to that for metal mining
(National Safety Council, 1980) which is considered a
rather hazardous occupation. Furthermore, the WPCF
surveys (Hadeed, 1980) indicate that not only is the
safety record poor, but that it has been declining in a
long-term trend since 1967.
24
-------
2. The WPCF survey indicates that the majority of accidents
happens to employees involved with collection and
distribution systems rather than the on-site treatment of
wastewater. The AWWA also sees a larger portion of
accidents off-site than on-site (Becker, 1981).
3. The WPCF data above indicate that in general the larger
wastewater treatment plants have poorer safety records.
This trend has been observed for some time for injury
frequency rates, but not necessarily the severity rates
over the same time frame (Hadeed, 1981).
4. Exposure to chlorine does not seem to be a major cause of
accidents. In the AWWA accident category above, exposure
to chlorine would fall in the category "contact with
radiations, caustics, toxic and noxious substances". This
comprises only 4 percent of the total accidents and puts
it on a par with insect bites. However, the actual risk
may be greater for chlorine exposure because of a greater
severity factor. In comparing the AWWA category data to
that for the State of Colorado, it appears that
wastewater treatment accidents are very similar to those
of all industries. The accidents caused by exposure to
chlorine (chemicals) is not significantly higher in water
treatment plants than in the general industry.
Risk Analysis Procedure—
The above data can be used to estimate the on-site accident
rate for chlorine as a disinfectant. The AWWA category "contact
with radiations, caustics, toxic and noxious substances" is
almost entirely due to exposure to chlorine(Becker 1981). A minor
contributor is exposure to organic solvents which are used in
cleaning equipment. A conservative estimate would indicate that 4
percent of accidents are caused by exposure to the disinfectant
chlorine.
The following procedure can be used to estimate the on-site
accident rate for chlorine as a disinfectant.
1. Use WPCF severity rate data based on plant size
2. Use AWWA 4 percent figure to assign lost workdays from
chlorine related accidents.
This simplistic procedure is based on three important
assumptions:
1. The drinking water treatment plant data from AWWA is
similar to that of wastewater treatment plant accident
rates.
25
-------
2. The severity (lost workdays) from exposure to
chlorine is the same as the average severity from all
types of accidents.
3. Chlorine accidents scale with plant size as all
accidents.
Transportation Hazards
Introduction—
In this section, accident data from the Department of
Transportation and U.S. census data on total chlorine movements
are combined to yield output of the form:
x(i)/y for railroad, truck (tank, 1 ton container or
cylinders), or barge,
where,
x(i) = deaths, injuries, dollars, or amount released;and
y = metric ton-kilometer.
This output, coupled with point-of-origin and point-of-
destination information, will enable the reader to estimate the
risk associated with the transportation of chlorine to a specific
wastewater treatment plant. Point-of-origin information can be
obtained from the map and tables of chlorine producers contained
in this section and the appendices. A sample risk calculation for
a particular treatment plant and details of the computations
which were used to generate tables are also in this section.
Chlorine is transported as solid calcium hypochlorite,
liquid sodium hypochlorite, or a compressed gas. However, almost
all the chlorine used for wastewater disinfectation is
transported as a compressed gas. For economic reasons, the
percentage of total chlorine shipped for wastewater disinfection
in the form of hypochlorite is neglible (Mitchell, R., Chlorine
Institute).
Data Base—
Department of Transportation.—Since 1971 there has been a
legal requirement that a report be filed with the Materials
Transportation Bureau of DOT for every accident involving the
transportation of a commodity. Since; January 1971, the reports
have included information on the number of deaths or injuries,
and the amount of property damage; and since January 1976, the
reports have also included the amount of material released. A
computer print-out was acquired from the Materials Transportation
Bureau covering the period from 1/71 - 12/80 (Morgan, 1980) and
was used as an accident data base. The data for railroad, truck,
26
-------
and barge are summarized in Table 6. Note that in this table, the
railroad accidents are shown with and without the major February
1978 rail accident in Youngstown, Florida.
TABLE 6. COMPILATION OF DOT ACCIDENT DATA, 1/71-12/80
Railroad
Railroad
Excl .
No. Accident
Reports
72
71
Deaths
8
0
Injuries
247
87
Property
Damage
$
1,111,498
22,498
Amount
Released
kg
1. 402x10-'
0.993xl05
Youngstown
Truck:
Cylinders
to 114 kg
0.911 Ton
containers
Tanker
Trucks
Barge
14
4
2
2
0
0
0
0
60
15
/I
3
8
23
15
,003
,550
,000
0
574
245
23
See
note
Note - Information not available prior to 1976.
Bureau of Census.—Total chlorine movements by
transportation mode and shipment weight can be obtained from the
Commodity Transportation Survey of the U.S. Bureau of the Census.
These data are compiled once every five years; the last published
data are for the year 1972 (U.S. Bureau of the Census, 1972).
Chlorine movements for 1972, in metric ton-kilometers, by
transportation mode and by shipment weight are listed in Table 7.
The breakdown of Census data by weight class and totals by
transportation mode allowed total truck movements (15.0%) to be
divided among tank trucks, 0.91 metric ton containers and 114 kg.
cylinders. The results of this calculation can be obtained in the
following manner. The 0.1% and 5% listed in Table 7 under 453 kg
and 454-4539 kg, respectively, can be assigned to 114 kg
cylinders and 0.91 metric ton cylinders, respectively. Tanker
trucks were then assumed to carry the remaining 9.9% of the truck
shipments. The result of this data analysis is shown in Table 8.
A continued analysis of this sort shows that the values for truck
shipments given in Table 8 are consistent with al 1 of the data in
Table 7.
The 1977 data was obtained prior to publication (R.Torene,
1981). Unfortunately, the unpublished 1977 data are less complete
than the published 1972 data. Indeed, the quality of the 1977
data, particularly that of shipment weight, precludes the kind of
analysis used to obtain Table 8. Since accident rates are
27
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TABLE 7. PERCENT DISTRIBUTION OF CHLORINE SHIPPED BY
TRANSPORTATION MODE AND BY SHIPMENT WEIGHT (TONMILES) FOR 1972
A. Transportation Mode
Rail
Truck (combines DOT motor carrier
and private truck data)
Water
84.9
B. Shipment Weight
Under 454 kg
454 - 4,539 kg
4,540 - 13,619 kg
13,620 - 27,239 kg
27,640 - 40,859 kg
40,860 kg and over
0.1
5.0
6.3
6.1
2.5
80.2
100.2
TABLE 8. BREAKDOWN OF CHLORINE SHIPMENTS
BY TRANSPORTATION MODE AND CONTAINER
Percentage
(for 1972)
Annual Average
metric ton-km
(for 1971-1980)
Rail
Truck:
Cylinders to
114 kg
0.91 Metric Ton
Cylinders
Tank Truck
Barge
84.9
0.1
5.0
9.9
0.3
100.2
1858
2.2
109
219
6.4
2194.6
calculated for truck shipments specific to container size, the
1972 data are used as a measure of the percentage of chlorine
shipped by transportation mode and container for the period 1971-
28
-------
1980. The average of the totals from the 1971 and 1977 census
data, 2191 mil lion ton-km, was used as a measure of the average
yearly ton-km of chlorine shipped during this period.
Methodology and Calculation of Accident Rates—
The DOT data base over the period 1/71 - 1/80 (Table 6) was
averaged to obtain an estimate of deaths, injuries, and property
damage per year. The data base for the period 1/76 - 1/80 (Table
6) was averaged to obtain an estimate of chlorine released per
year. The percentage breakdown by transportation mode and
container shown in Table 8 was then used to calculate a
corresponding breakdown by ton-km, assuming a yearly total of
2191 million ton-km. The results of this analysis, shown in
Table 8, were used to normalize the accident data per ton-km.
The accident rates calculated by this procedure are shown in
Table 9.
TABLE 9. ACCIDENT RATES PER METRIC TON-KM
Deaths
Railroad 4
Railroad
excluding
Youngstown
Truck:
Cylinders to
114 kg
0.91 Metric Ton
Containers
Tanker Truck
Barge
.3xlO-10
0
0
0
0
0
Injuries
1.
4.
2.
1 .
3.
4.
4xlO~8
7xlO~9
7xlO~6
4xlO~8
2xlO~8
7xlO~8
Property
Damage
$
e.oxio"5
1.2xlO~6
3.6xlO~4
2.15xlO~5
0.7xlO~5
0
Chlorine
Released
kg
3.3xlO~5
2.3xlO~5
1.2xlO~4
1.0xlO~!>
4.6xlO~8
See note
Note - information not available.
Table 9 shows that the accident rates for truck-transported
cylinders are consistently higher than the other categories
listed. This is likely due, in part, to the fact that a greater
number of cylinders are needed to carry a given amount of
chlorine. The deletion of the Youngstown accident from the
railroad totals does not greatly alter the relative ordering of
the accident rates in Table 9 except in the category of property
damage per ton-km.
Location of Chlorine Producers—
Figure 1 (Chlorine Institute, 1980) shows the location of
operating chlorine plants in the U.S. The chlorine producers and
29
-------
packaging plants are listed in the Appendices. These data can be
used to obtain point-of-origin information in order to calculate
specific accident rates to a particular wastewater treatment
plant.
Sample Risk Calculation—
The risk associated with the transportation of chlorine to
the Metropolitan Denver Sewage Disposal District No. 1 is used as
a sample calculation. The chlorine used by Metro No. 1 is
produced by National Lead of Salt Lake City, Utah (Puntenny,
1981). The chlorine is shipped the entire Salt Lake-Denver
distance of 853 km by railroad (usually in 50 metric ton cars).
Total chlorine usage was 374 metric tons in 1980 and 481 metric
tons in 1981 (Puntenny,1981). Using the accident rates per ton-km
in Table 9, the following risk factors are calculated for the
410,300 chlorine metric ton-km exposure of 1981:
Deaths 1.8 x 10~4 per year
Injuries 5.7 x 10~3 per year
Property damage $25.00 per year
Chlorine released 14 kg per year
Process Reaction Products Hazards
Total Residual Chlorine—
Formation.—Chlorine applied to water in its elemental or
hypochlorite form initially undergoes hydrolysis to form free
available chlorine consisting of molecular chlorine, hypochlorous
acid, and hypochlorite ion. The relative proportions of these
free chlorine forms is pH dependent. At the pH of most waters the
hypochlorous acid and hypochlorite ion will predominate.
In wastewater effluent and other types of waters, free
chlorine reacts readily with ammonia to form monochloramine and
dichloramine. The presence and concentrations of these combined
forms depend on many conditions; chiefly pH, temperature, and the
initial chlorine to ammonia-nitrogen ratio. Both free and
combined chlorine may be present simultaneously. Chlorinated
wastewater effluents and certain chlorinated industrial effluents
normally contain only combined chlorine forms, and at the normal
pH levels of such effluents the predominant species (and
disinfectant) is monochloramine.
The oxidative products formed in chlorinated seawater are,
for the most part, the same as in freshwater although the complex
nature of sea water influences the relative abundance and type of
chemical species found. The high chloride ion concentration in
sea water influences the amounts of hypochorite, chlorite and
30
-------
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chlorate ions produced. The formation of chloramines is expected
but concentration is variable and generally lower than in fresh
water. Chlorine may react with other halogens (iodide, fluoride,
bromide), but the principal reaction in seawater involves
formation of hypobromous acid and hypobromite ion from bromine
(Carpenter and Maculady, 1978). Bromamines, bromate, and bromine-
chloride complexes may also form.
As with fresh water the nature and amount of chemical
species found as products of chlorination of salt water will
vary with the site-specific parameters of pH, temperature,
reaction time, sunlight, and chemical composition of the water.
Persistence.—Because hypochlorous acid is a strong
oxidizing agent, its stability in natural water is very low,
especially at low pH. Hypochlorous acid rapidly oxidizes
inorganic compounds and rates of reaction with organic compounds
are generally slower. Generalizations on rates of loss are
difficult to make because these rates depend on sunlight, depth
of stream turbulence, temperature, pH, and type of reactants.
Monochloramine, being more stable, would be expected to
persist for hours to days compared to the minutes to hours
persistence for free chlorine. The bromine and bromamines formed
in seawater are slightly weaker oxidizing agents, but they are
less stable and and thus less persistent. The persistence of
chloro-organics produced in chlorination is discussed in their
respective sections below.
Use and Occurrence in the Environment.—Wastewater
disinfection operations are presently governed by state water
guality criteria in terms of meeting or exceeding specific levels
of total or fecal coliforms in effluents. For example, California
has adopted a bacteriological limitation based on the most
probable number (MPN) of total coliform organisms. Eighty percent
of the samples of effluent must contain an MPN less than 1000/100
ml (median of 240/100 ml) for coastal bathing waters, a median of
70/100 ml for shellfish growing areas, and a median of 23/100 ml
for confined waters where human contact is possible with a
dilution of at least 100 to 1. In order to achieve a total
coliform count of 23/100 ml consistently in a well designed
chlorination system, good quality secondary effluent would
receive a chlorine dose in the range of 10-15 mg/L. The residual
from such a system would be on the order of 2-4 mg/L. In a review
of some 60 plants, White (1975) revealed that many of the plants
operate in this range of dosage and residual. Systems that are
inefficient may require dosages up to 25 mg/L with resulting
residuals as high as 8 mg/L. Site-specific wastewater
characteristics will have a great influence on the amount of
chlorine needed to meet standards.
32
-------
Residuals as high as those discussed above have a
deleterious effect on aquatic life existing near the outfall. The
most recent (1976) EPA criterion to protect aquatic life has set
chlorine residuals at 0.2 mg/L (0.5 mg/L instantaneous maximum)
free residual chlorine (FRC) for 2 hr/day, and to 0.002 to 0.010
mg/L total residual chlorine (TRC) for salmonid fish and marine
and other fresh water organisms, respectively. Brungs (1976) has
recommended a single criterion of 0.003 mg/L TRC for continuously
exposed fresh water aquatic life.
Effects on Freshwater Vertebrates.—In order to determine
the toxicity of chlorine to aquatic life, several factors must be
considered. Chlorine is introduced at variable doses into
freshwater ecosystems that are affected by a wide range of
environmental conditions such as temperature, water quality, and
species composition. The toxicity of the various forms of
chlorine residuals to freshwater organisms is somewhat species
specific. The generalization that cold water species (salmonids)
are more sensitive to chlorine than warm water species may be an
oversimplification in light of recent studies that have shown
minnows and catfish to have median lethal concentration (LC-50)
values close to those of the salmonids. At best, toxicity may be
similar at genus level but generalizations above this level
(i.e., family) are tenuous. There are conflicting data concerning
toxicity with respect to the size of the fish. The differences
may be due to test methods. Data on the effects of chlorine
residuals on eggs, larvae and reproductive ability are also
limited.
Avoidance behavior in fish to chlorinated effluents has been
demonstrated in the laboratory and the field (Seegert, G.L. and
Bogardus, R.B., 1980). The fact that fish kills related to
chlorine are rare (Seegert, G.L. and Bogardus, R.B., 1980) and
that lack of species numbers and diversity has been observed in
rivers below waste treatment plants (Tsai, 1968, 1970), indicate
fish can avoid low levels of chlorine residuals that are
continuously discharged. The lowest concentrations avoided were
0.005, 0.035, and 0.050 mg/L for the mimic shiner, white bass,
and bullhead minnow, respectively.
The metabolic functions of poikilothermic animals are
directly tied to temperature and hence may be expected to
influence the toxicity of chemicals in such organisms. Most of
the studies to date indicate that the resistance of fish
continuously exposed to chlorine is inversely related to
temperature. The effects of temperature changes on toxicity are
less a factor during long exposures than during short exposures.
Water quality characteristics play an important role in
determining the responses of fish to chlorine. The pH of effluent
water affects the relative proportions of the various chlorine
species. The few studies done in this area suggest the role of pH
33
-------
in toxicity is probably related to the chemical species and not
any direct effect on the organism. Many toxicity studies do not
state water quality parameters and measurement of TRC does not
reflect the toxicity of test solutions.
Effects p_n Freshwater Invertebrates.—Since invertebrates
are part of the food chain in the aquatic ecosystem, the toxicity
of chlorine residuals to these species is important. As with
fish, chlorine residual toxicity varies greatly with species. The
water flea, Daphnia magna, seems to be the most sensitive species
with decreased reproduction at 0.002 mg/L and 100% mortality at
0.125 mg/L TRC (Brungs, 1973; Arthur, et al., 1974). The most
resistant invertebrate species studied was the Oligochaete worm
with an LC-50 of 91.0 mg/L (FRL) (Chung, S.L., 1960). More work
is needed on invertebrate toxicology in order to judge the
effects of the other water parameters previously discussed on
these organisms.
Effects on Plants.—Aquatic plants act not only as shelter
for fish and invertebrates, but also as substrates on which such
organisms live. They are a food source, and they enrich the
aquatic ecosystem by fixing carbon, thus increasing those foods
necessary for energy expenditure. They also produce free oxygen
required by all aerobic organisms. This fixation of carbon and
the production of oxygen is done by the process of
photosynthesis. The algae, Chlorel la pyrenoidosa, the most
sensitive species found, has been shown to experience a 50%
decrease in growth at 0.18 mg/L (FRC, CRC) and a 43% mortality at
0.6 mg/L TRC (Kott and Edlis, 1969). Most of the species studied
had decreased growth at a concentration of 2.0 mg/L TRC.
The literature dealing with the effects of TRC on freshwater
organisms is massive. The key elements of each reported study
reviewed in this study are summarized in tabular form in the
Appendices. This mass of reported data is also summarized in a
general format in Figures 2 and 3.
The reported data were first grouped according to the
consequences observed as a result of chlorine exposure. Those
consequences are avoidance behavior, mortality threshold, 50%
mortality, and 100% mortality. These consequences are defined as:
Avoidance - detection of TRC by the organism, increase in
environmental stress, altered behavior, depressed activity;
Mortality Threshold - increased mortality above that caused
by natural cycles and events;
50% Mortality - corresponds to LC-50 data; significant
mortality and stress; and
100% Mortality - absence of most aquatic organisms.
34
-------
CO
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O.I
T
100% MORTALITY
50% MORTALITY
MORTALITY THRESHOLD
LJ 0.001
0.01 O.I 1.0
INSTREAM RES. CL2, mg/l
10.0
FIGURE 2. RESPONSE OF SELECTED FRESHWATER ORGANISMS TO TOTAL
CHLORINE RESIDUAL
35
-------
HH
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2
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X 60
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0 30
20
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JJ 1 1 1 1 1 1 JUT
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—
—
—
—
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o
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—
—
—
—
—
—
—
10.0
1.0 O.I 0.01
TOTAL RESIDUAL CHLORINE mg/l
0.001
-------
Each group of data was then plotted as a function of the
reported chlorine residual. Figure 2 shows a sample plot for
selected freshwater organisms. In Figure 3 the chlorine residual
ranges containing many reported values are shown as heavy bars
for each of the four consequences observed. The thin lines
extending from each bar cover chlorine residual ranges with a few
data reported. The ends of the thin lines represent the extreme
values reported for each consequence. The overlap of the bars
clearly illustrates the variability in the reported results. For
example, the bulk of the data reported as mortality threshhold
lie in the upper ranges of the data reported showing 50% and 100%
mortality. In other words, an instream TRC of 1 mg/L might be at
the mortality threshhold; or it might cause a 50 to 100%
mortality.
To determine qualitative aquatic impacts using Figure 3,
first estimate the effluent residual chlorine concentration
(i.e.TRC). Then estimate effluent discharge (Qe = effluent
discharge) and design stream flow (Qs = stream flow above
discharge point) using the same units for both flow estimates.
Next calculate Qe/Qs. Enter the graph at the lower lefthand scale
using the estimated effluent TRC concentration. Move horizontally
to the calculated Qe/Qs ratio line (if necessary interpolate
between the lines shown in Figure 3). Then, move down vertically
and read the in-stream residual chlorine (TRC) concentration.
With the in-stream TRC concentration estimated then move up
vertically to determine the possible effects on freshwater
organisms.
Toxicity of_ Chlorine t£ Marine Vertebrates.—When comparing
mortality data for several species of fish it was found that 96
hour LC-50 values were generally between 0.032 and 0.5 mg/L TRC
(Gentile, 1974). A species of flounder was found to have 50%
mortality at 2.5-10.0 mg/L over short time spans (0.345 mg/L).
Although Atlantic Menhaden and Winter flounder eggs and larvae
were unaffected at 0.5 (3 min) to 10.0 mg/L (20 min), several
species, White Perch, Atlantic silverside, Fundulus heteroclitus,
and Trinectes maculatus, when exposed to TRC concentrations from
0.03 - 0.08 mg/L (10 min), showed avoidance behavior (Gentile,
1974) .
Toxicity of Chlorine to Marine Invertebrates.—The oyster
Crassostrea vir. had a 46% decrease in ciliary movement when
exposed to 0.2 mg/L TRC, and a pumping threshold at 1.0 mg/L
(Galtsoff, 1962). The mussel Mytilus edulis was found to have
unattached young at concentrations as low as 0.02-0.05 mg/L ard
the same species had 100% mortality at 1.0 mg/L over 15 days
(James, 1967). Copepods of various species tested at different
exposure times had mortalities from 22% to 90% at concentration
ranges of 0.25 to 10 mg/L (Dressel, 1971; Coughlon and Davis,
1976; McLean, 1973). Barnacles had LC-90-100 values about 1.0
mg/L. Barnacle nauplii did not grow at 1.0 mg/L and the lethal
37
-------
concentrations for 12-62% of the test populations were 0.25 to
1.0 mg/L. Blue crab were shown to be affected by 0.10 mg/L (96
hour LC-50) while share crab had the same 96 hour LC-50 at 1.4
mg/L. The 96 hour LC-50 for sand shrimp was found to be 0.09 mg/L
though the LC-55 for 10 minutes was 10.0 mg/L (Patrick and
McLean, 1970).
Toxicity of Chlorine t£ Marine Plants.—The toxic effects to
marine phytoplankton range from 50% decreased growth at 0.03
mg/L to 5-10 mg/L for Giant Kelp (Clendenning and North, 1959;
Morgan and Stress, 1969). Most of the species tested were found
to show 50% decreased growth between 0.10 to 1.0 mg/L chlorine.
Data Summaries.—The TRC data summaries in the Appendices
also include reported data in each of the areas discussed above.
Chloroform—
Formation, definition, sources, and levels in the
environment.—Chloroform (CHC13) is a low molecular weight,
volatile hydrocarbon formed during the chlorination of wastewater
as we 1 1 as cooling and drinking water. Jol ley, et al.
(1973,1974,1975) report tliat approximately one percent of the
chlorine applied to wastewater forms stable chloroorganics like
chloroform. Unlike other chloroorganics discussed in this report,
there are data on the levels of chloroform in the effluents of
wastewater treatment plants. Concentrations generally range from
5 to 20 ug/L (NAS, 1978). Bellar, et al. (1974) report that the
concentration of chloroform in wastewater prior to chlorination
is 7.1 ug/L; after chlorination, the effluent chloroform
concentration is 12.1 ug/L.
Chloroform enters surface waters (and thus drinking water)
from a variety of sources other than wastewater treatment plants,
including: 1) precipitation and 2) industry (paper mills, rubber
manufacturers and chemical companies) (NAS, 1978). Effluent
chloroform concentrations from industrial sources are usually
higher than wastewater levels. For example, the NAS (1978)
reported levels in paper mill effluents from 10 to 20,000 ug/L.
Effluents from rubber manufacturers in Louisville and Calvert
City, Kentucky ranged in concentration from 2,600 to 22,000 ug/L
(NAS, 1978).
In the Region V organics survey of drinking water of 83
cities in the United States, the median concentration of
chloroform was 20 ug/L; the maximum concentration was 366 ug/L
(EPA, 1975). In the NORS survey of 80 municipal water sci.r-.es,
chloroform concentrations ranged from 0.1 to 311 ug/L; j.he median
concentration was 21 ug/L. Mean concentrations in the atmosphere
ranged from 0.045 to 4.0 ug/cubic meters.
38
-------
Chloroform is formed in wastewater by the reaction of
chlorine with natural and synthetic organic substances under
alkaline conditions. There are numerous sources of organic
matter, including humic substances, plant material, synthetics,
and the end products of metabolic reactions of aquatic
microflora.
The rates of these reactions are dependent on certain
conditions, such as pH, temperature, and water quality.
Generally, chloroform (and other trihalomethanes) production in
chlorinated water is greatest at high temperatures (about 40° C)
and neutral to high pH values (7 to 11), while the presence of
ammonia reduces the formation of trihalomethanes.
Chloroform resists decomposition at ambient temperatures.
Prolonged exposure to sunlight with or without air results in
some decomposition, but the rate is not appreciable. Degradation
of chloroform in water is accelerated by aeration and the
presence of certain metals, such as iron (Hardie, 1964).
Bioconcentration of chloroform is fairly insignificant
according to several published studies. Person and McConnell
(1975) studied the levels of chloroform at various trophic levels
and found no significant bioconcentration. The U.S. EPA (1978)
reported a bioconcentration factor of 6 over 14 days in bluegills
and a tissue half-life of less than one day.
Effects of Chloroform on Freshwater and Marine Organisms.—
In freshwater fish several studies to determine the LC-50 have
been carried out. In the Bluegill, Bentley, et al. (1975) report
an LC-50 of 100 to 115 mg/L. Rainbow trout, on the other hand,
are more sensitive to chloroform. Bentley, et al. (1975) report
an LC-50 of 43.8 to 66.8 mg/L for adult rainbow trout. Clayberg
(1971) reported 100% mortality in the orange-spotted sunfish at
106.9 to 152.7 mg/L. Apparently, chloroform is not highly toxic
to freshwater fishes.
Data on the effects of chloroform to freshwater
invertebrates are limited. The U.S. EPA report (1978) indicates
that the mortality threshhold (i.e., concentration at which
mortality commences) for Daphnia magna is between 1.8 to 3.6
mg/L. Furthermore, the LC-50 for Daphnia magna was reported to be
28.9 mg/L.
Several marine species have been studied in regard to their
response to chloroform. Jones (1947) reported that anesthesia
occurred within 90 minutes in Threespine Sticklebacks exposed to
207.6 mg/L. In the Ninespine Stickleback, Jones reported
avoidance at 148.3 to 296.6 mg/L. The Pink Shrimp, Peneus
duorarum, a marine species, was studied by Bentley, et al.
(1975); they reported an LC-50 of 81.5 mg/L.
39
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Effects of Chloroform on Humans.—There are three routes of
entry in humans: 1) inhalation with absorption through the
lungs, 2) ingestion with absorption through the gastrointestinal
tract, and 3) dermal absorption. Pulmonary exposure to chloroform
produces a rapid rise in blood levels of chloroform. Equilibrium
is reached between blood and inspired air approximately 80 to 100
minutes from initial exposure (Lehman and Hasegaula, 1910).
Absorption by the gastrointestinal tract is approximately 100
percent efficient (Fry, et al., 1972).
In mammals, including humans, ingestion of chloroform is not
considered highly toxic. Ingestion of 30 to 100 ml of chloroform
resulted in gastrointestinal disturbances accompanied by delirium
and narcosis (vonOettingen, 1964). In humans, the mean lethal
dosage is approximately 44 grams (Grosselin, et al., 1976).
Once in the blood, chloroform distributes throughout the
body. The highest concentrations are found in peripheral nerves
(Cornish, 1975). Interestingly, chloroform can be transported
across the placenta. Fetal liver concentrations are higher than
maternal liver concentrations (EPA, 1979).
Chloroform is partially metabolized by the liver and kidneys
of mammals. Furthermore, it has been reported that chloroform is
converted to CC>2 in the lung, but only to a smal 1 degree. The
majority of absorbed chloroform is expired or excreted unchanged.
Studies using 14C-labelled chloroform in rats have shown after 18
hours 74% of the labelled chloroform was exhaled unchanged
(vonOettingen, 1964). Fry, et al. (1972) report that 96% of the
radioactively labelled chloroform given to adult humans was
exhaled unchanged 8 hours after administration. Chiou (1975)
reported that the half life of chloroform in the blood is 1.5
hours.
Chloroform mutagenicity was tested in bacterial systems. In
the Ames Salmonel la typhimurium test, chloroform did not prove
mutagenic in strains TA1535 and TA1538. In Escherichia coli K-12,
chloroform did not prove mutagenic (Uehleke, et al., 1976,1977).
Chloroform's teratogenic potential has been assessed in a
number of studies, although most studies investigated inhalation.
For example, Schwetz, et al. (1974) report that in rats 30 ppm
inhaled for 7 hours per day on days 6 - 15 of gestation resulted
in wavy ribs and delayed skull ossification, while inhalation of
100 ppm produced missing ribs, subcutaneous edema, imperforate
anus. nd delayed skull ossification. Thompson, et al. (1974)
report fetal toxicity in rats and rabbits exposed orally to
chloroform.
The International Agency for Research on Cancer (IARC)
evaluated and published a review of several experiments which
attempted to evaluate the potential carcinogenicity of
40
-------
chloroform. These experiments, including evaluations by the
National Cancer Institute, were performed in mice, rats, and dogs
with a variety of routes of administration. Doses ranged from 15
to 100 mg/kg/day and duration of exposure ranged from 8 weeks to
7.5 years (in dogs). The IARC concluded that chloroform is
carcinogenic in mice (liver) and rats (kidney), and that
chloroform presents a carcinogenic risk to man.
Epidemiologic surveys of the potential cancer risk of
chloroform have been made. The IARC reviewed the study of Bomski,
et al. (1967), but reported that the study was inadequate to draw
any conclusions because of the small numbers used and the short
follow-up. On the EPA's request, the National Research Council in
1978 reviewed 10 studies on the association between cancer and
trihalomethane consumption in drinking water (EPA, 1979). The
Council concluded that in most studies, the exposure and duration
levels were inferred, and that there were inadequate controls and
other invalidating factors. Thus, it is impossible to make a risk
evaluation of chloroform in humans from this data.
The risk of developing cancer from the consumption of
chlorofrom has been calculated by the National Academy of
Sciences. Using a linear, non-threshhold extrapolation from
animal data, the lifetime risk was estimated to be 1.5 x 10~7
to 17 x 10 per microgram of chloroform per liter. This means
that 1.5 to 17 cases of cancer wi 1 1 occur in a population of
10,000,000 if the drinking water contains 1 ug/L chloroform. The
EPA also calculated risk levels and found that if the level of
chloroform is 1.21 ug/L, 1 case of cancer can be expected to
occur in a population of one million. Recently, however, the EPA
modified its calculation of risk.
Estimates of risk depend on accurate estimates of exposure.
Both the EPA and NAS have estimated the annual exposure of humans
to chloroform under a variety of conditions. While these values
are equivocal, they may at least be used to estimate approximate
cancer rates. The problem lies in determining which exposure
model is most accurate and most typical. From Table 10, it is
clear that annual exposure (in mg/yr) is the same in adults and
children at each concentration exposure. For instance, at a
minimum concentration exposure (0.0001 mg/L), annual uptake in
adult men, adult women, and children is calculated to be about
0.037 to 0.088 mg/yr. At median concentration exposure, the
uptake is estimated to be about 7.6 to 18 mg/yr and at maximum
concentration exposure, 13 to 320 mg/yr. These estimates include
chloroform uptake from tap water, as well as water-based drinks.
In Table 11, also from the NAS study, the estimated exposure
to chloroform from all sources is given. At minimum concentration
exposure (minimum levels in drinking water), fluids provide a
minor portion of the total chloroform uptake; at typical and
maximum exposure levels, however, fluids provide the largest
41
-------
TABLE 10. CHLOROFORM UPTAKE FROM FLUIDS(NAS,1978)
mg/year (assuming 100% absorption)
Exposure
Level
mg/L
0.0001
(min.
level)
0.021
(median
level)
Fluid
Source
Tap
water
All
Tap
water
All
Adult
Min.
Intake
0.
0.
3.
7.
016
037
44
57
Man
Max.
Intake
0.
0.
5.
18
027
088
59
.4
Adult
Min.
Intake
0.
0.
3.
7.
016
037
44
67
Woman Chi Id (5-14 yrs.)
Max. Min. Max.
Intake Intake Intake
0.
0.
5.
18
027
088 0.036
59
.4 7.67
0.061
12.8
proportion of the chloroform. From these data, it would appear
that the range in total annual chloroform exposure is quite large
— from a minimum of about 0.6 mg/yr to well over 800 mg/yr. If
this range is accurate, estimates of the cancer risk can be made
to indicate the extremes. Furthermore, the relative contribution
of drinking water to the cancer risk can be estimated.
TABLE 11. HUMAN UPTAKE OF CHLOROFORM
FROM ENVIRONMENTAL SOURCES (NAS, 1978a)
mg/year
Exposure
Level
Minimum
Typical
Maximum
Source
Fluid Intake
Atmosphere
Food
Totals
Fluid Intake
Atmosphere
Food
" Totals
Fluid Intake
Atmosphere
Food
Adult Man
0.037
0.41
0.21
0.66
14.90
5.20
2.17
22.27
321
474
16.4
Adult Woman
0.037
0.37
0.21
0.62
10.70
4.70
2.17
17.57
321
434
16.4
Child
0.036
0.27
0.21
0.52
10.70
3.40
2.17
16.27
223
310
16.4
Totals 811
771
549
42
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Table 12 shows estimates of uptake of chloroform from a
variety of sources, as determined by the EPA. Adult uptake varies
from a low of 3.14 mg/yr to a high of 563 mg/yr, a range less
than that determined by the NAS.
TABLE 12. HUMAN UPTAKE OF CHLOROFORM
FROM AIR, WATER AND FOOD (USEPA,1978b)
ExposureSourceUptakePercent of
Level mg/year Total Uptake
Maximum Atmosphere 204 3"6
Water 343 61
Food 16 3
Totals 563 100
Mean Atmosphere 20 21
Water 64 69
Food 9 10
Totals 93 100
Minimum Atmosphere 0.41 13
Water 0.73 23
Food 2.00 64
Totals 3.14 100
Trichloroethylene—
Formation, Definition, Sources and Levels in the
Environment.—Trichloroethylene (1,1,2-trichloroethylene, C2HCl3>
is a volatile chlorohydrocarbon formed during wastewater
disinfection as well as drinking water chlorination (Bellar, et
al., 1974). Little data is available on the production or
occurrence of trichloroethylene in wastewater treatment plants.
Bellar, et al. (1974) report that wastewater treatment plant
influent may contain as much as 40 ug/1 of trichloroethy lene.
This study also reported that chlorination increased the
trichloroethylene level from 8.6 ug/1 to 9.8 ug/1. Certainly,
more work is necessary to evaluate the production of
trichloroethylene from wastewater treatments plants.
Trichloroethylene also enters surface waters (and therefore
drinking water) from other sources such as precipitation and
industry (IARC,1979).
The U.S. National Organics Monitoring Survey (EPA,1976)
found that about one-fifth of the municipal water supplies tested
did not contain trichloroethylene and that the detected levels
43
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were quite low (EPA, 1978b). Mean concentrations of
trichloroethylene ranged between 1.3 to 21 ug/L (IARC, 1979).
Toxicity of_ Trichloroethy lene to Freshwater and Marine
Organisms—Trichloroethylene is toxic to freshwater fish; but,
like tetrachloroethylene, its toxicity is low (Alexander, et al.,
1978; Pearson and McConnell, 1975 and EPA, 1978b). The 96-hour
LC-50 for the Fathead Minnow is 36.5 mg/L with static testing and
40.7 with flow-through testing (Alexander, et al., 1978). The 96-
hour LC-50 for Bluegills is 44.7 mg/L under static testing
conditions (Alexander, et al., 1978).
Little work has been directed toward the elucidation of the
effect of trichloroethylene on freshwater invertebrates. Studies
on Daphnia magna indicate that the 48-hour LC-50 is approximately
85.2 mg/L under flow-through testing conditions (EPA, 1978b). In
life cycle tests, 10 mg/L had no effects (Alexander, et al.,
1978). The unicellular alga, Phaedactylum tricoinitum, showed a
50 percent decrease in the uptake of radioactive carbon during
photosynthesis at a concentration of 8.0 mg/L (McConnell and
Pearson, 1975). The Dab, a saltwater fish, had a 96-hour LC-50 of
16 mg/L (McConnell and Pearson, 1975). In the Cheepshead Minnow,
exposure to 2.0 mg/L of trichloroethylene produced erratic
swimming, uncontrolled movement and loss of equilibrium.
Salt water invertebrates have also been tested, but in a
rather limited manner. Grass Shrimp displayed erratic swimming,
uncontrolled movement and loss of equilibrium after exposure to
2.0 mg/L of trichloroethylene (Borthwick, 1977). In barnacle
nauplii the 48-hour LC-50 was 20 mg/L (McConnell and Pearson,
1975) .
On the whole, there are not much data concerning the
toxicity of trichloroethylene in aquatic ecosystems, which makes
the construction of dose-response and risk assessments
impossible. From the limited data above, however, it appears that
the toxicity of trichloroethylene is low. Furthermore, because
the levels of trichloroethylene are low in the effluents of
wastewater treatment plants, acute toxic reactions in fish and
other aquatic organisms below wastewater outfalls are unlikely.
Further studies may show that low levels of trichloroethylene
have chronic effects such as teratogenicity, mutagenicity and
carcinogenicity.
Effects on Humans.—Studies of the effects of
trichloroethylene on humans are limited. Most studies are of
occupationally exposed individuals who were subjected to rather
high concentrations of the chemical in the air. Most of these
studies indicate an effect on the central nervous system in
workers exposed to trichloroethylene; reported findings include
mild fatigue, decrease in psychomotor function, decrease in
44
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performance ability, and headache (Stuart, et al., 1970; Stopps
and McLaughlin, 1976; Salvini, 1971; Nomiyama and Nomiyama,
1971).
No data were available on the teratogenicity of
trichloroethylene in humans, and animal studies show little, if
any, effect. Schwetz, et al. (1975) tested mice and rats for
possible teratogenic effects, but their results were
inconclusive.
No human data are available for assessment of the mutagenic
potential of trichloroethylene. Mutation rate was increased by
the administration of trichloroethylene to a number of bacterial
strains including E_._ coli and Salmonella typhimurium (IARC,
1979). Similar findings were obtained in the yeast, Saccharomyces
cerevisiae; however, impurities in the technical grade
trichloroethylene used in the these experiments now have been
shown to be mutagenic agents (IARC, 1979).
Only one study of the possible carcinogenic potential of
trichloroethylene in humans was found. The incidence of cancer
mortality was studied in a population of 518 workers
occupationally exposed to low levels of trichloroethylene.
Exposure was detected by urine analysis. No excess in cancer
mortality was noted in this study; however, the limited size of
the study group precluded detection of cancers such as liver
cancer which are found in a low incidence (Axelson, et al., 1978
in IARC, 1979).
Studies in laboratory animals indicate that
trichloroethylene may be carcinogenic. For instance,
trichloroethylene induces transmutation in Fisher rat embryo
cells in vitro, a study system used for the identification of
carcinogens. Second, a 1976 NCI study indicated an increase in
the incidence of hepatocellular carcinoma in mice; Osborne Mendel
rats, however, showed no increase in cancer (IARC, 1979).
Tetrachloroethylene—
Formation, definition, sources and levels in the
environment.—Tetrachloroethylene (C2C14) is a volatile
chlorohydrocarbon (also known as perchloroethylene) formed in
wastewater disinfection (Bellar, et al., 1974). Unfortunately,
there is little data on the level of tetrachloroethylene found in
wastewater treatment plant effluents. Bellar, et al. (1974)
reported that chlorination of wastewater slightly increases the
concentration of tetrachloroethylene with final effluent
concentrations of ,oout 4.2 ug/L.
Tetrachloroethylene enters surface waters (and thus drinking
water) from a variety of industrial sources other than wastewater
treatment plants (Shakelford and Keith, 1970).
45
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Generally, levels of tetrachloroethylene are quite low in
surface waters. The Safe Drinking Water Commission (1977) found
that 8 of 10 samples from different water supplies contained
tetrachloroethylene, but only in low concentrations, ranging from
0.07 to 0.46 ug/L (IARC/1979). In the U.S. EPA's National
Organics Monitoring Survey (EPA,1976), tetrachloroethylene was
detected in 9 of 105 drinking waters sampled. Again, levels were
quite low; for instance, the mean concentration of the nine
samples was 0.81 ug/L and the range was 0.2 to 31 ug/L (U.S. EPA,
1978) .
Effects of_ Tetrachloroethylene on Aquatic Organisms.—
Tetrachloroethylene is toxic to freshwater fish, but the toxicity
is relatively low compared to total residual chlorine. In the
Bluegill, the 96-hour LC-50 is approximately 12.9 mg/L (EPA,
1978). In the Fathead Minnow, the 96-hour LC-50 is approximately
18.4 to 21.4 mg/L (Alexander, et al., 1978). The Bluegill and
Fathead Minnow are both warm water species. Unfortunately, no LC-
50 data are available for cold water species. The 96-hour LC-50
for the Sheepshead Minnow was 29.4 to 52.2 mg/L (EPA, 1978),
slightly higher than the LC-50 for freshwater fishes.
Freshwater invertebrates are also affected by
tetrachloroethylene. Unfortunately, evidence is sparse. The 48-
hour LC-50 for Daphnia magna is 17.7 mg/L (EPA, 1978).
Unicellular organisms such as algae are affected by
tetrachlorethylene; however, in these organisms its toxicity is
relatively low. In several species of algae, EC(50) (effective
concentration) ranged from 10,000 to over 800,000 mg/L (Pearson
and McConnell, 1975; EPA, 1978).
In summary, because of the lack of data on the toxicity of
tetrachloroethylene in fishes, a dose-response curve cannot be
constructed. It does appear evident though, at least for the
species discussed above, that since tetrachloroethylene
concentrations in the effluents of wastewater treatment plants
are low, acute toxic reactions in stream fish below the outfall
are unlikely. Further studies may show that low levels have
chronic effects such as mutagenicity, carcinogenicity, and
teratogenicity.
Effects of Tetrachloroethylene on Humans.—The chief effects
of tetrachloroethylene in humans are central nervous system
depression, hepatotoxicity and neural disorders. Most studies of
these effects, hov^ver, examine the inhalation of fairly high
concentration? of tetrachloroethylene.
Only two systematic studies of human carcinogenicity were
found. These studies indicate that tetrachloroethylene may be
carcinogenic. Blair, et al. (1978) mentioned a clinical report of
five cases of chronic symptomatic leukemia in a family that
46
-------
operated a dry cleaning business. Blair, et al. (1979) studied
death certificates from deceased laundry and dry cleaning
employees who had been exposed to tetrachloroethylene,
trichloroethylene and chloroform. He observed an excess of cancer
in certain loci: cervix, lungs and skin. Liver cancer and
leukemia exhibited a slight excess. Blair's study was
inconclusive due to the small sample size.
Animal carcinogenicity studies show mixed results. B6C3 mice
orally exposed to large doses of tetrachloroethylene for 78 weeks
(5 days per week) displayed an increase in hepatic carcinoma
(NCI, 1977). Rats, similarly exposed, did not have an increase in
cancer (IARC, 1979).
The mutagenicity of tetrachloroethylene, like its
carcinogenicity, is not resolved. Cerna and Kypenova (1977)
tested Salmonel la typhimurium (TA 100) and found some
mutagenicity; however, in mice tetrachloroethylene did not appear
to be mutagenic (Cerna and Kypenova, 1977). Greim, et al. (1975)
failed to elicit a mutagenic response in E._ coli treated with
tetrachloroethylene.
Chlorophenols—
Formation, Definition, Sources and Levels in the
Environment.—The Chlorophenols are nonvolatile
chlorohydrocarbons which are formed during the chlorination of
wastewater as well as drinking water (Jolley, 1973; Jolley, et
al., 1975; Glaze and Henderson, 1975). Chlorine reacts readily
with phenol in aqueous solutions over a wide range of pH values
(Carlson and Caple, 1977). Levels in the effluents from
wastewater chlorination are in the .001 mg/L range.
Burtschell, et al. (1959) proposed a scheme for the
mechanism of phenol chlorination in water. Monochlorophenols (2-
and 4-chlorophenol) are formed first; further chlorination of
these species produces 2,6- and 2,4-dichlorophenol. After 18
hours of reaction, the products consisted of: 1) less than 5
percent monochlorophenol, 2) 25 percent 2,6-dichlorophenol, 3) 20
percent 2,4-dichl'orophenol and 4) 40 to 50 percent 2,4,6-
trichlorophenol.
Like other chlorohydrocarbons discussed in this report,
Chlorophenols enter surface waters from a variety of sources
other than wastewater treatment facilities. Industrial processes
are the dominant source of these pollutants (EPA, 1976; IJC,
1980) .
Effects of Chlorophenols on Aquatic Organisms.—A
substantial amount of data have been reported on the effects of
Chlorophenols on aquatic organisms. Key elements of these reports
are summarized in tabular form in the Appendices.
47
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The toxicity of the chlorinated phenols to aquatic organisms
varies with the degree of ring substitution as well as the
position of the chlorine atoms on the ring. In general, toxicity
increases with increasing substitution. The more toxic
multichlorinated products may come from industrial effluents
rather than wastewater and drinking water chlorination (IJC,
1980) .
Among the freshwater fish, Bluegills (Lepomis macrochirus)
are slightly more sensitive than Fathead Minnows (Pimephales
promelas), Goldfish (Carassius auratus) or Guppies (Poecillia
reticulata). LC-50 (96 hour) for Bluegills range from 0.140 to
10.0 mg/L (Table 13).
In Fathead Minnows toxicity (96-hour LC-50) varies,
depending on the degree of substitution, from 0.03 to 20.17 mg/L.
Forty-eight hour LC-50's in Rainbow Trout (Salmo gairdnerii)
range from 0.023 to 10.0 mg/L.
The toxicity of chlorophenols to aquatic invertebrates has
been fairly well documented (Jolley, 1973; Jolley, et al., 1975;
Glaze and Henderson, 19V:>). In Daphnia magna, toxicity ranges
from 0.29 to 7.43 mg/L; again, the variation in toxicity
(measured as 96-hour LC-50) results from the degree of
substitution. Altogether toxicity among fishes and invertebrates
(i.e., Daphnia) is fairly similar. The response of freshwater
algae and duckweed to a variety of chlorophenols has been studied
fairly extensively. There is a wide range in the effective
concentrations resulting from differences in chemical species as
well as plant species. It appears that polychlorinated phenols
are most toxic (Huang and Gloyna, 1967 and 1968; EPA, 1978;
Blackman, 1955; Erickson and Freeman, 1979).
Limited data are available on the toxicity of chlorophenols
to marine fishes and invertebrates. Several studies of the.
Sheepshead Minnow show that toxicity (measured by the 96-hour LC-
50) range from 1.66 to 5.35 mg/L (EPA,1978).
Mysid shrimp appear to be less sensitive to chlorinated
phenols. A considerable amount of data is available regarding the
effects of chlorinated phenols on saltwater algae. Again
depending on the end point under consideration, chemical species,
and algal species, the 96-hour LC-50 are present in a wide range
from 0.25 to approximately 8 mg/L (EPA, 1978; Erickson and
Freeman, 1979).
In summary, it would appear that chlorophenols are toj.ic,
but not as much as residual chlorine. If the available data are
representative of wastewater disinfection facilities, then stream
concentrations would be expected to fall well below toxic levels.
48
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However, significant bioconcentration can result in toxicity
which might not be predictable from the acute toxicity data
previously reviewed in this section. Bioconcentration factors in
fish for various chlorophenols range from 8 to 240 with the more
substituted compounds having higher factors (EPA, 1980).
Effects of Chlorophenols in Humans.—No data could be found
on the levels of chlorophenols in ambient air. Toxicity of
chlorophenols in mammals is well known. 2-chlorophenol, as well
as 3- and 4-chlorophenol, is an "uncoupler" of oxidative
phosphorylation (Mitsuda, et al., 1963). In addition, these
compounds are known to cause convulsions (Farquharson, et al.,
1958; Angel and Roger, 1972). Apparently, gastric acidity allows
chlorophenols to be converted to more soluble forms which are
readily absorbed by the intestines.
In rats, 2-chlorophenol produces symptoms including
restlessness, increased respiration rate, followed by weakness,
tremors, convulsions, dyspnea, coma and death (Farquharson, et
al., 1958). At high concentrations, 2-chlorophenol produced fatty
degeneration of the liver, renal granular dystrophy, and
degeneration of intestinal mucosa (Bubnov, et al., 1969).
Acute poisoning with 2,4,5-trichlorophenol impairs physical
activity by inducing motor weakness and convulsion; however, the
trichlorophenols tend to produce fewer and milder convulsions
than the monochlorophenols (Farquharson, et al., 1958; Deichmann,
1943). Weinbeck and Garbus (1965), Parker (1958) and Mitsuda, et
al. (1963) report that, like the monochlorophenols,
trichlorophenols uncouple oxidative phosphorylation, but only
weakly.
A number of reports have disucssed the carcinogenicity of
the chlorophenols. Table 13 summarizes the results of work
reported in this area.
A study of tumor promotions in mice by chlorophenols was
done by Boutwell and Bosch (1959). 20% solutions were applied
weekly to female Sutter mice (2-3 months old) for 15 weeks. This
followed the application of a tumor initiator, DMBA (9,10-
dimethyl -1,2-benzanthracene). Two monochlorophenols (2- and 3-)
were found to promote papillomas. 4-chlorophenol was not tested.
The trichlorophenols (2,4,6- and 2,4,5-) were tested in the same
manner with 2,4,5- increasing the incidence of papillomas in DMBA
pretreated mice. A bioassay was conducted by the NCI on 2,4,6-
trichlorophenol in the feed of F344 rats and B6C3F1 mice. The NCI
concluded from these studies that 2,4,6-trichlorophenol was
carcinogenic in male rats (F344) (producing lymphomas or
leukemias). In addition, this compound was also found to be
carcinogenic in both sexes of B6C3F1 mice (producing
49
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hepatocel lular carcinomas and adenomas). It remains to be seen at
this writing, whether 4-chlorophenol, 2,3,6-trichlorophenol, and
2,3,4,6—tetrachlorophenol are carcinogenic.
TABLE 13. CARCINOGENIC ASSESSMENT OF CHLOROPHENOLS
IN LABORATORY SPECIES
Compound Results
2-chlorophenolpossible promoter
3-chlorophenol possible promoter
4-chlorophenol not known
2,3,6-trichlorophenol not known
2,4,5-trichlorophenol possible promoter
2,4,6-trichlorophenol possible carcinogen in
rats and mice
2,3,4,5-tetrachlorophenol not known
The EPA has calculated cancer risk level for only one of
the chlorophenols, 2,4,6-trichlorophenol. The probability is 0.5
that a 2,4,6-trichlorophenol concentration of 3.6 ug/1 will cause
1 additional case of cancer in a population of 1 million people.
Monochlorophenols have not been tested for mutagenicity. Two
test systems have been used to test the mutagenicity of
trichlorophenols. Fahrig, et al. (1978) found that 400 mg of
2,4,6-trichlorophenol increased the mutation rate in
Saccharomyces cerevisiae. Rasanen, et al. (1977) reported that
several polychlorinated phenols (2,3,5-, 2,3,4,6-, 2,3,6-, 2,4,5-
and 2,4,6-) were not mutagenic when tested via the Ames test.
Dichlorobenzenes—
Formation, Definition, Sources and Levels in the
Environment.—The dichlorobenzenes (DCBs) are chlorinated
benzenes produced during the chlorination of wastewater (Glaze
and Henderson, 1975). Three isomers of dichlorobenzene exist:
1,2-dichlorobenzene, 1,3-dichlorobenzene , and 1,4-
dichlorobenzene. The DCBs are readily soluble in fats (Windholz,
1976) and are relatively volatile (Jordan, 1954; Kirk and Othmer,
1963; Varschueren, 1977). Glaze and Henderson (1975) reported
that levels of DCBs in effluents from wastewater treatment plants
were about 10 ug/L.
DCBs have been detected in rivers, groundwater, industrial
and municipal wastewater, air, and soil. Ware and Ware (1977)
reported that the average levels of 1,2-DCB in industrial
wastewater were 2 mg/L. Glaze, et al. (1976) reported that DCB
can be produced during the chlorination of municipal wastewater.
50
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Glaze and Henderson (1975) reported that both 1,2-DCB amd 1,4-DCB
concentrations were about 10 ug/L in chlorinated municipal
wastewater effluents.
The EPA (1975) reported detecting low concentrations of DCBs
in municipal drinking water samples. For 1,2-DCB, 1,3-DCB and
1,4-DCB reported concentrations were 1, <3, and 1 ug/L,
respectively-
Effects of Dich 1 orobenzene on Aquatic Organisms.—Dawson
(1977) found the 96-hour LC-50 for 1,2-DCB in Bluegills was 27.0
mg/L, while the EPA reported 96-hour LC-50s of 4.28 and 5.59
mg/L. The large difference in the reported LC-50 values may
result from different methods of chemical dispersion used by
investigators. In the Fathead Minnow, the EPA reported the 96-
hour LC-50 for 1,4-DCB to be 4.0 mg/L. Rainbow trout were
slightly more sensitive than Bluegills and Fathead Minnows
according to the 1980 EPA study in which LC-50s for 1,2-DCB and
1,4-DCB were reported as 1.58 and 1.12 mg/L, respectively. The
EPA (1980) performed analyses of,the toxicity of 1,2- and 1,4-DCB
on Fathead Minnows during the embryo-larvae stage (ELS). Results
of this work indicate that this may be the most sensitive stage.
Values for ELS studies range from 0.56 and 2.5 mg/L (EPA, 1980).
Only a few studies have been performed to examine the toxicity of
DCB on freshwater invertebrates. The 48-hour LC-50s for 1,2-DCB
and 1,4-DCB were 2.44 and 11.0 mg/L, respectively, in Daphnia
magna (EPA, 1978). The Midge was less responsive to these
pollutants with a 48-hour LC-50 of 11.76 and 13.0 mg/L for 1,2-
DCB and 1,4-DCB, respectively (EPA, 1978). Studies in algae show
that reduction in chlorophyll a. and reduction in cell number
occur at roughly 90 to 100 mg/L (EPA, 1978).
Very little data exist on the toxicity of DCBs to saltwater
organisms. From the existing data, it appears that shrimp are
more sensitive than fish and plant species. Two species of
saltwater fish have been tested: the Tidewater Silverside and
the Sheepshead Minnow. The 96-hour LC-50 for 1,2-DCB was 7.3 mg/L
for the Tidewater Silverside. The 96-hour LC-50's for 1,2-DCB and
1,4-DCB in the Sheepshead Minnow are 7.66 and 7.44 mg/L,
respectively. Thus, it appears that saltwater fishes are less
sensitive to the DCBs than freshwater fishes.
In Mysid Shrimp, the 96-hour LC-50s for 1,2-DCB and 1,4-DCB
were 1.97 and 1.94 mg/L, respectively.
In algae, 96-hour ECSO's (measurements 01 50 percent
reduction in chlorophyll and cell number) range between 44 and 59
mg/L.
Bioconcentration of DCBs may, however, cause chronic toxic
reactions. In Bluegills, the steady-state bioconcentration
factors (BCF) for the 1,2-DCB, 1,3-DCB and 1,4-DCB are 89, 66 and
51
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60 , respectively (EPA, 1978). Neely, et al. (1974) calculated
the steady-state BCF in Rainbow Trout; his calculations show a
BCF of 210.
In summary, reported wa'stewater effluent DCB levels are so
small relative to reported DCB toxicity levels that instream
toxic effects from DCB in effluents are unlikely to occur.
Effects of DCB's on Humans.—A number of studies report the
presence of DCBs in human tissues. For instance, West and Ware
(1977) detected DCBs in human blood in residents of New Orleans.
Morita, et al. (1975) report the presence of 1,4-DCB in human
adipose tissue.
Most of the clinical cases of DCB poisoning reported have
been through inhalation (16 of 22) (Girard, et al., 1969; Burners,
et al., 1952; Weller and Crellin, 1953; Perrin, 1941; Cotter,
1953; Gadrat, et al., 1962; Petit and Champeir, 1948; Nalbandian
and Pierce, 1965; Campbell and Davidson, 1970; Downing, 1939;
Frank and Cohen, 1961; Ware and West, 1977). The quantitative
efficiency of absorption via inhalation has not been determined.
Three of the above 22 cases of poisoning resulted from
ingestion but no quantitative absorption efficiency has been
determined via this route (Campbell and Davidson, 1970; Frank and
Cohen, 1961; Hallowell, 1959). Animal experiments indicate that
GI absorption of DCBs is rapid since effects, excretion and
metabolites have been observed within 1 day of oral exposure
(Rimington and Ziegler, 1963; Azouz, et al., 1953; Poland, et
al., 1971).
Human and animal studies have indicated absorption via
dermal exposure. Three of the 22 clinical cases involved dermal
exposure (Girard, et al., 1969; Dowing, 1939; Nalbandian and
Pierce, 1965). The dermal application of 1,2-DCB (5 times twice
daily applications) to the abdominal skin of rats caused dermal
absorption (West and Ware, 1977).
Because of low water solubility and high lipid solubility,
DCB's should be able to cross barrier membranes (West and Ware,
1977). This would allow the DCB's to be widely distributed in
various tissues. Lipid soluble halobenzenes may accumulate in the
body and reach toxic levels and also may recirculate for long
periods (West and Ware, 1977). The clinical and experimental data
indicate wide distribution resulting in changes in blood and
blood chemistry, neuromuscular function, a--d liver and kidney
structure and function.
Numerous studies of acute and chronic toxicity of DCBs have
been reported in humans as well as laboratory animals. Because of
their wide circulation in the body, DCBs affect numerous
processes. They do not appear to be site specific like other
52
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chemicals discussed in this report (e.g., chlorophenols). All of
the available literature are either follow-up studies of human
ingestion or inhalation studies; and, in all cases, exposure
levels do not correspond to the low level, chronic exposure found
in drinking water.
Recent work using mixed isomers and 1,2-DCB failed to
demonstrate carcinogenic response (EPA, 1980; Varshavskaya,
1967b). Earlier work, however, seems to indicate that 1,4-DCB may
increase the rate of leukemia in rats and may promote cancer in
irradiated mice (Murphy and Sturm, 1943; Parsons, 1942).
Anderson, et al. (1972) reported that 1,2-DCB was not
mutagenic in the Ames test. However, several authors have
reported chromosomal breakage and cell abnormalities in plants
exposed to DCBs (Carey and McDonough, 1943; Sharma and
Bhattacharyya, 1956; Sharma and Sarkar ,1957).
5-Chlorouracil—
Among the volatile chloroorganics found in the chlorinated
effluents of wastewater treatment facilities is 5-Chlorouracil
(5-CU) (Jolly, 1973, 1974, 1975; Jolley, et al., 1976; Glaze and
Henderson, 1975). Because of the possible incorporation of 5-CU
into nucleic acids and the potential mutagenicity and/or
carcinogencity, this compound was selected for review of its
known toxicity.
Effects on Aquatic Organisms.—There are only three
available reports on the toxicity of 5-CU on carp eggs and
embryos. There is some disagreement in the results of these
studies which renders their interpretation difficult. Eyman, et
al. (1975) studied the effects of 5-CU on embryos one hour after
hatching and demonstrated increased mortality (1.2 to 4.3%) at
concentrations ranging from 0.5 to 10.0 mg/L. Gehr, et al. (1974)
demonstrated reduced hatching success in carp eggs exposed to 7.0
mg/L 5-CU. In contrast, Trubalka and Burch (1978 and 1979)
studied carp embryos and were unable to observe any effects of 4-
7 day exposures to 5-CU concentrations ranging from 0.01 to 100
mg/L.
Two studies on the effects of 5-CU in Daphnia magna have
appeared in the published literature. Gehrs, et al. (1972)
reported delayed and reduced production of offspring in Daphnia
exposed for 7 days to 0.01 mg/L 5-CU; while Gehrs and Southworth
(1976) reported no change in median survival time in Daphnia
exposed to 5-CU in concentrations ranging from 0.01 to 100 mg/L.
Clearly, one can tentatively conclude that in Daphnia
reproduction is impaired at levels considerably below mortality;
thus, reproductive impairment could be the factor which limits or
decreases population size if concentrations reach critical
levels.
53
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Johnson, et al. (1977) carried out extensive studies on the
incidence of mortality in a marine species, the Spotted Sea
Trout. They studied eggs and larvae exposed to a variety of
concentrations. The results of their work are shown in Table 14.
These results show that 2-hour eggs are generally more sensitive
than 10-hour eggs. Larvae are less sensitive than 2-hour eggs at
10 and 100 mg/L; at all other concentrations larvae display
either similar or slightly higher mortality than 2-hour eggs.
TABLE 14. EFFECTS OF 5-CHLOROURACIL
ON SPOTTED TROUT EGGS AND LARVAE
Concentration Percent Mortality
mg/L 2 hr. eggs 10 hr. eggs 1 hr. larvae
100
10
5
1
0.5
0.1
0.01
100.0
59.0
7.5
5.5
5.0
2.0
0.5
44.5
24.5
12.0
11.5
17.5
4.0
3.0
56.5
30.0
10.0
9.0
6.4
2.0
0.5
Note - exposure temperature = 25U C
test type, static, concentrations not measured
test duration = 48 hours
Gehrs and Southworth (1976) studied the toxicity of 5-CU, 4-
chlororesorcinal and a complex mixture of chloroorganics. Their
results show possible antagonistic effects on Daphnia.
Effects of 5-CU on Humans. — 5-CU was found to be mutagenic
in E._ coli WP-1; however, 1 gram/liter of 5-CU in the drinking
water of mice for more than a year produced no observable genetic
or somatic effects (Gumming, 1976). This concentration approaches
the maximum water solubility and is about 1,000,000 times greater
than the estimated environmental level. Gumming (1976) reported
that 5-CU was incorporated into the DNA in the livers and testes
of mice but did not examine these or other sites for genetic
damage. Gumming (1976) tested for dominant lethal mutations in
male mice mated with unexposed females, but this test showed no
significant results. Gumming (1976) used the specific locus test
to measure recessive visible mutations at seven loci in mice; in
this study, no mutations were observed.
The carcinogenicity and teratogenicity of 5-CU are not
known.
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Summary of Chlorination Byproduct Toxicity.
Table 15 summarizes reported wastewater effluent levels and
measured effects ranges for each of the chlorination products
discussed above. Several important conclusions can be
drawn from these data.
1. For all byproducts except TRC, effluent levels are
below levels known to be acutely toxic. Thus, any
stream dilution would reduce concentrations below
those reported to have acute toxic effects.
2. Effluent levels of TRC are within the concentrations
known to have acute toxic effects. Thus, even with
stream dilution TRC will be expected to have acute
toxic effects.
TABLE 15. EFFLUENT LEVELS AND EFFECTS RANGES
FOR SELECTED WASTEWATER CHLORINATION BYPRODUCTS
Byproduct Reported Effluent Measured Effects
Levels Range
(mg/L) (mg/L)
Total Residual
Chlorine 1-8 0.001-10
Chloroform 0.012-0.020 1.0-300
Tetrachloro-
ethylene 0.004 10-800
Trichloro-
ethylene 0.01-0.04 1.0-80
Chlorophenols 0.0005-0.03 0.01-500
Dichlorobenzenes 0.01 1.0-200
5-chlorouracil 0.004 0.01-100
Note - Upperlimits of rangesindicate upperlimits of testing.
Please note, the effects ranges in Table 15 include only acute
toxic effects - mostly LC-50 data. These data do not include
long-term chronic effects in aquatic organisms. Because of
bioaccumulation, seemingly low levels of a chemical in effluents
and receiving stream water may have biological significance.
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CHLORINATION/DECHLORINATION
INTRODUCTION.
An alternative to releasing a chlorinated wastewater
effluent with a high residual chlorine (TRC) level which may be
toxic to aquatic life is to dechlorinate or to remove the
residual chlorine. The process of chlorination/dechlorination
thus involves chlorination to disinfect the secondary wastewater
followed by a dechlorination step to remove or reduce the
residual chlorine remaining from the disinfection step. Although
dechlorination has been used successfully in a variety of
drinking water situations, it has only recently been investigated
for wastewater.
There are several common methods of dechlorination:
physical adsorption using activated carbon; chemical reduction
using sulfur dioxide (S02)/ sodium sulfite (Na2S03), or other
sulfur containing reducing agent; photochemical decomposition
using ultraviolet radiation, (Seegert, 1978) and holding ponds
for the dissipation of chlorine residuals (Can, et al., 1979). Of
these methods, the use of sulfur dioxide is the most cost
effective and appears to have the greatest promise for wastewater
applications (White, 1972) (Can, et al., 1979). Reasons for
sulfur dioxide popularity for dechlorination are centered around
its similarity to chlorine — commercial availability as a
compressed liquid, transportation modes, handling of liquified
gas, and safety precautions — as well as cost effectiveness. Due
to similar physical and chemical properties for chlorine and
sulfur dioxide, equipment is often interchangeable. Therefore,
this risk analysis concentrated on sulfur dioxide as a
dechlorinating agent.
HAZARD IDENTIFICATION.
The identification of hazards, or risks, associated with the
chlorination/dechlorination wastewater disinfection process
should include the risks within the production, transportation
and handling, and use categories for both chlorine and sulfur
dioxide. In addition, there may be identifiable risks from the
combined use of the two chemical agents which may be absent when
each is analyzed separately. The risks associated with the use of
chlorine are treated above. Production risks were omitted from
this study in the scope of work.
Sulfur dioxidv. i-; used in numerous industrial applications
— as a refrigerant, a bleaching agent, a disinfectant, a liquid
solvent, and as a raw material for the production of suIfuric
acid (Shreve, 1967). Thus, its present and potential use as a
dechlorinating agent represents a small fraction of the sulfur
dioxide produced in the United States. Neglecting the risk
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associated with the production of sulfur dioxide for the purpose
of dechlorination will have a negligible effect on the overall
risk associated with the production of sulfur dioxide.
The risks associated with sulfur dioxide will be similar to
those for chlorine due to similar chemical and physical
properties, and transportation and handling methods.
The primary risks associated with the transportation and
handling of sulfur dioxide are:
1. human exposure to liquid sulfur dioxide;
2. human exposure, both occupational and public, to gaseous
sulfur dioxide; and
3. vegetation exposure to either liquid or gaseous sulfur
dioxide.
Human exposure to liquid sulfur dioxide can cause serious
skin or eye burns; however sulfur dioxide vaporizes at -10°C at
atmospheric pressure (Handbook of Chemistry and Physics, 1969).
Thus, most exposures will be to gaseous rather than liquid sulfur
dioxide.
Sulfur dioxide is a colorless gas with a suffocating odor.
The principal human exposure routes are inhalation followed by
eye and skin contact. The taste threshold is approximately 0.3 to
1 ppm, and the odor threshold is approximately 0.5 to 1 ppm (U.S.
Department of Health, Education, and Welfare, 1970, and Faith,
1972). Sensitive individuals notice sulfur dioxide in the 1 to 2
ppm range while most people show respiratory irritation when
chronically exposed to 5 ppm concentrations. Severe bronchospasms
can be initiated in the 5 to 10 ppm range (Faith, 1972). When
sulfur dioxide contacts atmospheric moisture, sulfurous acid
(^503) may be formed which is a toxic irritant and is also
highly corrosive.
Different plant species vary considerably in their
susceptibility to exposure to gaseous sulfur dioxide. The
response to sulfur dioxide is also dictated by factors such as
temperature, humidity, plant age, and soil moisture. Acute
exposures to sulfur dioxide cause leaf discoloration and
destruction. Threshold concentrations start at about 1.0 ppm for
a one hour exposure (alfalfa). Chronic exposures result in
yellowing of leaf tissue, leaf drop, lesions, and suppression of
growth. Chronic symptoms occur at much lower concentrations,
typically less than 0.03 ppm. Low concentrations of sulfur
dioxide can also react synergistically with ozone to cause injury
to some sensitive plant species (Faith, 1972).
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The primary risks associated with the use of chlorine/sulfur
dioxide as a mechanism of disinfection and dechlorination are:
1. human and aquatic life exposure to residual chlorine,
2. human and aquatic life exposure to chlorine reaction
products,
3. human and aquatic life exposure to sulfur dioxide
reaction products,
4. human and aquatic life exposure to both sulfur dioxide
and chlorine reaction products,
5. aquatic exposure to low dissolved oxygen effluent, and
6. aquatic exposure to low pH.
The risks associated with residual chlorine and chlorine
reaction products are discussed above.
When sulfui- dioxide is dissolved in water, sulfurous acid
(H2S03) is formed, which is the actual dechlorinating agent.
Sulfurous acid reacts virtually instantaneously with both free
and combined chlorine to form chlorides, sulfates, and bisulfates
as the primary reaction products (White, 1972). Sulfurous acid is
also a moderately strong acid which will dissociate to form
bisulfite (HSCU~) and sulfite (S0o2~). Research is required to
identify the list of inorganic and organic reation products
associated with the use of sulfur dioxide as a dechlorination
agent. Research is also required to determine what effect, if
any, sulfur dioxide has on the chlorine reaction products formed
in the chlorination step or those which previously existed in the
wastewater effluent prior to chlorination. Nevertheless, the
primary products from dechlorination should be inorganic salts
such as chlorides, sulfates, and sulfites which have been shown
to be nontoxic to aquatic life at concentration levels usually
associated with wastewater treatment (Can, 1979).
The use of a chemical dechlorination agent like sulfur
dioxide requires an accurate and fail-safe injection and
monitoring system (Seegert, 1978). The addition of too little
sulfur dioxide can lead to a chlorine residual remaining in the
wastewater effluent which can be toxic to aquatic life at very
low levels. The monitoring of the residual chlorine before
dechlorination is the usual method of varying the r:qi^ired sulfur
dioxide (called feed forward control). For feedback control, the
limit of detection for residual chlorine analysis may be above
the maximum recommended residual chlorine level required for the
protection of aquatic organisms. The addition of too much sulfur
dioxide can result in injury to aquatic life. To prevent low
dissolved oxygen, an aeration process may be required downstream
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from the dechlorination step (U.S. Environmental Protection
Agency, 1976). However, Gan et al.(1979) found that reaeration is
not necessary, because there was negligible reduction in
dissolved oxygen as a result of sulfur dioxide dechlorination.
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SECTION 5
HAZARD IDENTIFICATION FOR OZONE
Ozone may be considered one of the most promising
alternatives to chlorine as a wastewater disinfectant. This
section deals with the identification and quantification of the
hazards associated with ozone as a wastewater disinfectant.
The identification of hazards, or risks, associated with
ozone is performed by considering the production, transportation
and/or handling, and use cycles in a manner similar to that for
chlorine. The instability of ozone dictates that it be produced
on-site; thus, primary risks associated with ozone can be limited
to on-site production and use. Knowledge of treatment plant
facilities for the production and use of ozone allows detailed
hazard identification.
The quantification of risks associated with ozone is much
more difficult than the quantification process for chlorine.
Although ozone has been extensively used as a disinfectant for
drinking water treatment in Europe, there is very limited use of
ozone for wastewater treatment. Furthermore, in the United States
chlorine is almost universally used for disinfecting water and
wastewater, and ozone's limited use in the United States has been
for taste, odor, and color control in the treatment of water
(Layton, 1972). The use of ozone as a wastewater disinfectant is
expanding in the United States (Venosa, 1972; Hais and Venosa,
1978), but experience is moderate.
The primary risks associated with the use of ozone as a
wastewater disinfectant are:
1. Human and vegetation exposure to gaseous ozone;
2. Human and aquatic exposure to residual ozone;
3. Human and aquatic exposure to ozonation reaction products
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ON-SITE USE HAZARDS
The primary on-site risks associated with the production of
ozone are:
1. human exposure, both occupational and public, to gaseous
ozone;
2. vegetation exposure to gaseous ozone in the vicinity of
wastewater treatment facility; and
3. human contact with high electrical voltage.
Human exposure to gaseous ozone is associated with
headaches, irritation of respiratory tract, and irritation to
eyes with changes in visual acuity. For sensitive individuals,
irritation from ozone starts near the odor threshold, 0.02 to
0.05 ppm by volume. Current human exposure standards for ozone
are 0.1 ppm for a period not to exceed 8 hours. This requirement
dictates efficient treatment facility design to collect and
destroy exhaust ozone. One plant has experienced ambient
concentrations of ozone in the 15 to 30 ppm range during initial
start-up conditions (Rakness and Hegg, 1979). Elevated ozone
concentrations can also be injurious to vegetation in the
vicinity of the treatment facility.
Ozone is partially soluble in water, and it is difficult to
obtain concentrations of more than a few milligrams per liter in
aqueous solutions under normal conditions of pH, temperature, and
pressure. Thus, at high applied dosages as might be used in the
disinfection of industrial and municipal wastewater, the
atmosphere above the contact tanks will be rich in both oxygen
(Op) and ozone (03). If the ozone concentration in the off gas is
not reduced by mechanical or chemical means, the discharge offgas
remains rich in ozone. Ozone is 1.5 times as dense as oxygen and
has a long ha If-life in the ambient atmosphere of approximately
12 hours (Miller, et al., 1978). Consequently, there is a real
possibility of high atmospheric concentrations of ozone in the
vicinity of the wastewater treatment facility, which can be
injurious to both plant and animal life.
The limited data available on the use of ozone for
wastewater disinfection prevent a quantitative assessment of the
accident rate for workers in wastewater treatment facilities
that use ozone. Furthermore, although the Europeans have
considerable experience using ozone as a disinfectant for potable
water, there are no parallel data for accident statistics. In
fact, ozone in Europe is not seen as a particular cause of
accidents or injuries (Bres, 1981).
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PROCESS REACTION PRODUCT HAZARDS
The hazards associated with ozonation reaction products must
be considered due to the variety and toxicity of reaction
products associated with the use of chlorine. Although the
consensus in the literature (Hais and Venosa, 1978; Winklehaus,
1977) indicates that ozone produces fewer harmful intermediate
and end products than does chlorine, the fact remains that very
little work has been done to identify ozone reaction mechanisms
and to identify ozone reaction products under field conditions.
Ozone can react by several mechanisms in wastewater
depending on pH, presence of metals, and types of organic
compounds present (Winklehaus, 1977). In general, oxidation
reactions of organic compounds with ozone are not complete
(formation of H20 and C02) thus leading to the production of both
stable and unstable intermediate and end product compounds.
Smaller, saturated organics are less reactive (more refractory)
to ozonation. Ozone products identified under field and
laboratory conditions include low molecular weight alkanes,
aldehydes, organic acids, and heterocyclics (Chappell, et al.,
unpublished; Kuo, et al., 1977). Although ozonation end products
may accumulate after repeated cycles of water reuse and thus
constitute a risk, there is little known about ozonation products
and their toxicity.
Effects of Ozone Residual on Aquatic Organisms
Hubbs (1930) investigated several water purification
processes for treating water supplies for fish. One of the
processes studied was ozonation. The responses of several fish
species included altered locomotive and respiratory movements,
followed by loss of equilibrium and wild swimming with
alternating quiescent periods when the fish rested on its side or
back. The quiescent periods lengthened and finally terminated in
death. Fish could recover from the symptoms of altered locomotive
and respiratory movements but not from lost equilibrium even when
they were transferred to a suitable water supply. Residual ozone
concentrations as low as 0.09 mg/L were lethal to fish in flow-
through assays. Crayfish died after exposure to 1.16 mg/L
residual ozone and unspecified planktonic and bottom
invertebrates were killed by residual ozone concentrations of
1.25 mg/L. Oxygen super-saturation, pH, and C02 content were
eliminated as possible causes of mortality. The methods used by
Hubbs to measure the ozone residuals (unbuffered potassium
iodide) may have led to errors i.r reported levels (Schechter,
1973).
Giese and Christensen (1954) studied the effects of residual
ozone in static tests on freshwater protozoa and rotifers. The
organisms were studied in minute volumes of water under a
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microscope, and measurements of ozone residuals were not done.
The residuals were probably low but the concentrations were high
enough to kill the organisms.
Giese and Christensen (1954) studied sea urchin
(Strongylocentratus purparatus) and marine worm (Urechis caupo)
eggs in unmeasured ozone concentrations and found changes in the
membranes and cortex of the animals.
McLean, et al. (1973b) studied the effects of ozonated
seawater on the eggs of the commercial American oyster
(Crassostrea virginica). The residual ozone concentration was
estimated to be less than 0.20 mg/L in a static bioassay. Eggs
spawned in the ozonated seawater exhibited an increase in
fertilization defects (decreased polyspermy and parthenogenesis)
and intercellular abnormalities (retarded meiosis and cleavage,
irregular polar bodies and abnormal nuclear cleavage) compared to
eggs spawned in unozonated seawater.
A residual ozone concentration of 0.01 to 0.06 mg/L was
demonstrated to cause 100% mortality within four hours to
Rainbow Trout. But when the same lakewater for hatchery use was
aerated for 11 minutes prior to delivery to the trout tanks the
mortality was eliminated (Rosenlund, 1975).
Barnacles (Balanus sp.) which were continuously exposed to
ozonated seawater (with a residual of 0.4 to 1.0 mg/L) died after
several days (up to a week) exposure. (Mangum and Mcllhenny,
1975)
Exposure to a residual ozone concentration of 0.1 mg/L for 5
minutes was lethal to marine phytoplankton (Skeletonema lostatum,
Chlorel la sp., Nannochloris sp., and Monochrysis lutheri) within
24 hours. Crab zoea exposed for one minute to 0.08 mg/L of
residual ozone showed a 0-20 percent mortality after 24 hours and
30-40 percent mortality after 48 hours. A one minute exposure to
0.2 mg/L residual ozone resulted in 100 percent mortality within
24 hours for crab megalops. Atlantic Silverside (Menidia menidia)
exposed to 0.08-0.2 mg/L residual ozone were killed within 30
minutes (Toner and Brooks, 1975).
Acute and chronic flow-through laboratory bioassays were
conducted with domestic secondary wastewater effluent disinfected
by ozonation and diluted by lake water. Seven fish species (Brook
trout, Coho salmon, Fathead minnow (fish and eggs), Whitesucker,
Walleye, Yellow perch, Largemouth bass) and six invertebrates
(Amphipods, Stonefly, Caddisfly, Crayfish, Operculate snail, and
Pulmonate snail) were tested for seven days in acute bioassays.
Two generations of one species of fish (Fathead minnow) and two
invertebrate species (Daphnia magna, and the amphipod, Gammarus
pseudo1imnaeus) were assayed in a chronic test. A residual ozone
concentration of 1-2 mg/L disappeared so rapidly from the bottom
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of the ozone contact column to the test-tanks that no measureable
levels were detected in the test water for both tests. No
measureable toxicity to aquatic life was found from either long-
or short-term exposures. When procedures were adjusted to shorten
the retention time so that a residual ozone concentration could
be maintained continuously in the effluent, an acute exposure to
0.2-0.3 mg/L was lethal to Fathead minnows after 1-3 hours
(Arthur, et al., 1975).
Increased hatchability and improved survival was reported
for fry from Rainbow trout eggs incubated in recirculating
aquarium water treated with ozone and deozonated with activated
carbon to a residual below 0.10 mg/L (Benoit and Matlin, 1966).
Rosenkranz, et al. (1978) reported that the 96-hour LC50
for adult White perch (Morone americana) exposed to ozone-
produced oxidants (OPO) was 0.22 mg/L. A significant decrease in
blood pH and an increase in hematocrit over the test period (96
hours) were observed in fish exposed to 0.10-0.15 mg OPO/L.
Histological changes in the gills of fish exposed to 0.01-0.15 mg
OPO/L over 24 hours were seen, but gill repair was evident after
a 14 day recovery period.
Richardson, et al. (1978) exposed 12-hour old Striped bass
(Morone saxatilis) eggs to a range of 0.005-0.10 mg OPO/L, then
evaluated the effects on egg development at 24 and 42 hours after
fertilization and the effects on two prolarval stages (24 and 48-
hour post hatch). The results indicated that OPO were more toxic
in freshwater after exposures in both fresh and estuarian water.
The eggs were more resistant than the prolarvae. Delayed hatching
occurred in eggs exposed to 0.05 and 0.10 mg/L OPO (estuarine
water), and the delayed hatch organisms survived better than the
hatched larvae at the same stage of development. Wedemayer, et
al. (1979) found that the acute toxicity curve for dissolved
ozone yielded a 96-hour LC50 of 0.093 mg/L O3 for juvenile
Rainbow trout. The authors reported death due to acute exposure
was most likely due to severe gill lamellar epithelial tissue
destruction accompanied by massive hydro-mineral imbalances.
Chronic tests showed little damage at 0.002 mg/L and some
histopathological changes in gill tissue at 5 ug/L O3.
The 24-hour LC50 for bluegill (Leponis macrochirus) was
0.06 mg/L and periodic dosing of six 30-minute exposures, 8 hours
apart, gave an LC50 of 0.32 mg/L. Bluegills held in
concentrations between 0.01 and 0.02 mg/L showed irregular
respiration, decreased activity, cessation of feeding, and random
mortality during a 6-week exposure (Paller and Heidinger, 1979).
Ward, et al. (1976) conducted bioassays using ozonated
wastewater obtained from a Michigan treatment plant that receives
mostly domestic wastes. Long-term life cycle toxicity tests were
conducted with several cold and warm water fish species and one
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invertebrate species. Fathead minnows were exposed to filtered
ozonated wastewater diluted from zero to 100% with wellwater and
with an 03 residual from 0.00 to 0.016 mg/L 63 for 30 days (first
generation) and 60 days (second generation). The first generation
of fish exhibited no lethal effects attributable to residual
ozone. The second generation of fish reared in about the same
residual ozone concentration (0-0.013 mg/L) and dilutions as the
first generation exhibited no definite lethal response. The
authors suggest that long-term exposure to ozonated effluent
would not be toxic or lethal to Fathead minnows. In the same
study no significant differences in length or weight were
observed over the life cycles of two generations of Fathead
minnows. The same species of fish reared in various
concentrations (0-0.016 mg/L) and dilutions (0-100%) of ozonated
effluent showed no adverse reproductive effects. There appeared
to be no adverse effects on the hatchability of eggs produced and
incubated in water with a ozone residual of 0.00 to 0.01 mg/L.
In experiments where 14 species of cold and warm water fish
were exposed to 100% ozonated effluent within 10 minutes after
disinfection, goldfish and fathead minnows survived an ozone
residual of 0.047 to 0.185 mg/L for 7-15 days. Under similar
conditions lake trout finger lings suffered 100% mortality after
five hours at an ozone residual of 0.322 mg/L. The other species
tested showed no mortality to ozone concentrations of 0.002 to
0.38 mg/L but due to ozone generation problems some of these
species were not tested for 96-hours.
Ward, et al. (1976) found Daphnia magna (less than 24 hours
old) had a 30% mortality when exposed to 100% ozonated effluent
(residual O3 0.03 mg/L) for 96 hours.
Ward, et al. (1977) used the same methods as in their
previous study (Ward, et al., 1976) to examine the acute and
chronic toxic effects on several freshwater fish and
invertebrates. The effluent to be ozonated was from a secondary
treatment plant in Wyoming, Michigan that receives 35-45% of its
waste from light industry (metal plating plants, dairy products)
and 55-65% from domestic sources.
The results of the life cycle test were much the same as in
the previous study. Adult fathead minnows exposed to residual
ozone concentrations at 0.003 to 0.01 mg/L showed no mortality
due to the ozone. Fathead minnow fry exposed to 0.001 mg/L 0-j
over 60 days produced poor survival data that was inconclusive
due to poor water quality not associated with the ozone
disinfection.
Ward (1976) reported that residual ozone in ozonated
secondary effluent diminishes to a concentration less than 0.01
mg/L within about 15 minutes after dosing. Arthur, et al. (1975)
also found in their toxicity studies with the fathead minnow that
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an ozone dose rate of about 5.7 +- 1 mg/L was needed to disinfect
and attain a residual of 1 mg/L . At an ozone residual level of
10 mg/L dissipation of ozone in the effluent was so rapid that
levels in the test tanks were unmeasureable and non-toxic.
Rosenlund (1975) also reported that water with a residual of 0.01
to 0.06 mg/L became nontoxic when aerated for 11 minutes before
introduction to fish.
Bioassays to determine ozone levels and doses to prevent
unwanted introduction of fishes into a reservoir were conducted
by Coler and Asbury (1980a). Larvae and, in some species, eggs
were exposed to ozonated lake water in static and continuous
tests. The 24-hour LC50 values ranged from 4.0 mg/L for channel
catfish eggs to 0.19-0.31 mg/L for rainbow trout larvae. In
another study by the same authors (1980) larvae and, in some
fish, eggs were assayed in a flow-through system at different
time periods and different concentrations to determine LC50
values for residual ozone exposures. In both of the above studies
eggs were more tolerant than larvae to ozone concentrations.
o_f the Effects of_ Residual Ozone
Residual ozone is toxic to aquatic organisms even at low
concentrations. The general pattern of toxicity for freshwater
fish appears to begin at a concentration of about 0.001 mg/L with
gill damage and loss of equilibrium. These conditions may result
in the subsequent death of some fish but unless the concentration
reaches the 0.01 mg/L range, where 50% to 100% mortality may
occur, most fish should recover with the dissipation of the
ozone.
The larvae of most species of fish seem to be more sensitive
to ozone than fish eggs. Adverse effects appear in the 0.1 mg/L
range for larvae and 1 mg/L range for eggs. The tolerance to
ozone by fish eggs appears to be associated with the membrane.
Benoit and Mat 1 in (1966) indicated that ozone may act on the
membrane only resulting in death if the membrane or gelatinous
matrix is disrupted causing lysis or leakage. Since adult or fry
fish will experience adverse effects at a concentration at least
one order of magnitude lower than the effect levels for eggs and
larvae, the early life stages of fish should be protected if
residual ozone levels are maintained below 0.001 mg/L range.
Freshwater invertebrates seem to be less sensitive to
residual ozone than fish. Mortality for most species does not
begin unti. 'he concentration reaches 0.01 mg/L range.
Laboratory estimates of ozone decay rates in Connecticut
River water at 9 and 22 degrees Celsius showed a linear
logarithmic decrease from 2.0 to 0.05 mg/L in 30 and 60 minutes,
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respectively (Coler and Asbury, 1980). Thus, when ozonated
effluents are aerated and then diluted rapidly, concentrations of
ozone will dissipate below the toxic levels.
Some studies indicated that ozonated water may be beneficial
to the growth and survival of fish eggs and fry after the
residual has been removed or dissipated (Benoit and Mat 1 in, 1966,
Ward, et al., 1976). The benefit may be due to the increase in
the dissolved oxygen (DO) content of ozonated water.
Supersaturation of oxygen in water does not seem to have great
toxic effects on fish at long exposure times (Wolke, et al.,
1974; Boack, et al., 1976; Nebeker, et al., 1979).
Although the data indicate that residual ozone is toxic to
aquatic life at low concentrations, it is not very stable in
water and will likely dissipate to concentrations below the toxic
levels before any adverse effects on aquatic life are noted.
Ozone residuals are typically measured at the exit side of the
contact basin which may be some distance upstream from where the
disinfected wastewater effluent is discharged into the receiving
body of water. In addition, the residence time for dissolved
ozone in aqueous solutions is very short, since even in a
nonreactive environment of distilled water, the half-life of
ozone is only 20 to 30 minutes (Miller, 1978).
Ozone remaining in effluents of ozone treatment plants will
dissipate long before any such water would be incorporated into
the drinking water supply system. Thus, the human health risk
from ozone residuals in wastewater effluents is insignificant.
Effects of Ozone By-products on Aquatic Organisms
The by-products of wastewater ozonation are dependent on the
organics (the precursors) present in the wastewater prior to
ozonation, and the number of potential precursors is very large.
The ozonated effluent of the Upper Thompson Sanitation District
treatment plant in Estes Park, Colorado, was analyzed for organic
residuals by Chappell, et. al. (1980); and a large number of
simple aliphatic and aromatic compounds were found. In order to
develop a sense for the hazards associated with the ozonation by-
products, the organic compounds discussed below were selected for
research in the literature.
n-heptane. — Several studies were cited in Vershlueran (1977)
regarding n-heptane toxicity. In Mosquito Fish a concentration of
5,600 mg/L had no apparent effect whereas toxicity was reported
at 1,000 mg/L. The 24-hour LC-50 in Mosquito Fish was reported to
be 4,900 mg/L. In Goldfish the 24-hour LD-50 was reported at 4
mg/L. In young Coho Salmon no significant mortality was induced
at concentrations below 100 mg/L after 96 hours in artificial
seawater at 8 C. These data obviously demonstrate a wide range of
toxicity.
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n-octane.—Verschlueran (1977) reported only one study of
toxicity for n-octane. In this study young Coho Salmon exhibited
no mortality in concentrations of less than 100 mg/L when exposed
for 96 hours.
n-hexanal.—No data available on aquatic toxicity.
m-xylene.—Verschlueran (1977) found the 24-hour LC-50 for
Goldfish to be 16 mg/L. Hann and Jensen (1974) reported that the
LC-50 was 10 to 100 mg/L. A few studies on the toxicity of o-
xylene were available. In Goldfish, the 24-hour LC-50 for o-
xylene is 13 mg/L (Birge, 1979). Brenneman, et al. (1976) report
the 96-hour LC50 for o-xylene in Goldfish at 14 mg/L. Hann and
Jensen (1974) reported that the 96-hour LC50 for Fathead Minnows
was 42 mg/L. In young Coho Salmon, o-xylene produced no
significant mortality in fish exposed 24 to 96 hours to 10 ppm;
the 24-hour LC-50 was determined to be 100 mg/L. The alga,
Chiorel la vulgaris, exposed for 1 day to 55 mg/L, suffered a 50
percent decline in cell number. Walsh, et al. (1977) reported a
96-hour LC-50 for xylene (no isomer given) in Rainbow Trout at
13.5 mg/L.
n-heptanal.—No toxicity data available.
n-nonanal.—No toxicity data available.
Ozone by-products would be produced in such small quantities
(ppt) (Chappell, et. al., 1980) in wastewater plant effluents
that the human exposure risks are negligible. Furthermore,
dilution of these products by receiving streams results in even
less risk. Finally, the toxic dose-response levels for ozone by-
products are orders of magnitudes greater than the expected
effluent levels so adverse effects to aquatic organisms are not
1ikely to occur.
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SECTION 6
HAZARD IDENTIFICATION FOR ULTRAVIOLET RADIATION
The use of ultraviolet radiation to disinfect wastewater is
now receiving attention as a possible alternative to the use of
chlorine. To become a viable alternative to chlorine in the
disinfection of wastewater, ultraviolet radiation must be safe
and effective.
The identification of hazards, or risks, associated with
ultraviolet radiation is performed by considering the production,
transportation and/or handling, and use cycles. Ultraviolet
radiation is produced on-site, which eliminates the
transportation and handling problems. Risks associated with the
use of ultraviolet radiation as a wastewater disinfectant are
human and aquatic exposure to insufficiently disinfected
wastewater effluent and human and aquatic exposure to ultraviolet
radiation reaction products. Primary risks associated with
ultraviolet radiation are limited to on-site production and use.
The quantification of risks associated with the production
and use of ultraviolet radiation is very difficult when only
considering published or historical information. Although there
has been some effort in using ultraviolet radiation for the
sterilization of potable water and the bactericidal effects of
ultraviolet radiation have been known for many years, there has
been very little application of ultraviolet disinfection of
secondary wastewater on any scale. Most of the potable water
sterilization systems using ultraviolet radiation have been small
in scale and drawing a parallel comparison between wastewater and
potable water again will be difficult.
ON-SITE USE HAZARDS
The primary on-site risks associated with the production of
ultraviolet radiation are:
1. human exposure to ultraviolet radiation
2. human exposure to ozone
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3. human contact with high electrical voltages.
Although direct exposure to ultraviolet radiation is
improbable in a well designed wastewater treatment facility,
accidental exposures to treatment plant personnel would carry
significant risk to human health. Human exposure to ultraviolet
radiation primarily affects the skin and eyes. Reddening, or
burning, of the skin due to ultraviolet radiation exposure is
most pronounced at a wavelength of 260 nm which is nearly
identical to the major radiation wavelength, 254 nm, produced by
the mercury discharge lamps used in disinfection. The severity of
skin burning is a function of the total ultraviolet dosage and
ranges from simple reddening to blistering and peeling of skin
with possible severe secondary effects. Data indicate that
ultraviolet radiation may produce or initiate carcinogenesis in
human skin. Ultraviolet exposure to eyes can result in damaged
corneas, impairment of visual acuity, and eye fatigue. With the
exception of cornea damage, most effects are temporary (HEW,
1973) .
Injurious threshold levels for ultraviolet radiation depend
on the particular wavelength. To prevent reddening ol unprotected
skin the American Medical Association has published an exposure
limit of 0.5 uWatt/square centimeter for exposure at 254 nm for
up to seven hours. The American Conference of Governmental
Industrial Hygenists has proposed a limit of 1,000 uWatt/square
centimeter for 300 to 400 nm radiation for a period of 16 minutes
to protect both skin and eyes. Since average ultraviolet
disinfection dosages for wastewater are in the range of 10,000 to
100,000 uWatt-sec/square centimeter, accidental exposure to
ultraviolet radiation poses a human health risk.
Ultraviolet radiation is capable of producing ozone when
oxygen, or air, is irradiated with low wavelength (<200 nm)
ultraviolet radiation. When the UV lamps and quartz sleeves are
maintained at the optimum temperature, air can be circulated
around the lamps, and thus the exhaust air would be rich in ozone
(Sheible, 1979). Waste treatment plant personnel could therefore
be exposed to ozone if the exhaust air is not properly treated or
vented. Ozone production from ultraviolet exposure of dissolved
oxygen in the wastewater appears to be insignificant. This is
primarily due to the low level of dissolved oxygen in wastewater
and the absorbing nature of the wastewater thereby diminishing
the ultraviolet radiation levels after a short distance from the
lamp source.
RADIATION AND REACTION PRODUCT HAZARDS
Ultraviolet radiation is classified as a physical
disinfecting agent. The primary germicidal effects of ultraviolet
radiation are due to the absorption of UV radiation by the
genetic material (deoxyribonucleic acid (DNA) and ribonucleic
70
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acid (RNA)) of the microorganisms. The resulting dimerization of
pyrimidine bases in the genetic material distorts the molecule
and prevents proper replication of cell material. This results in
death, mutation of offspring, and inviable offspring (Sheible and
Bassell, 1979). Photoreactivation is defined as restoration of
ultraviolet lesions in a biological system with light of
wavelength longer than that of the damaging radiation (Johnson,
et al.; 1979). Thus, when ultraviolet irradiated wastewater
effluent is exposed to radiation in the visible region, a repair
mechanism may be activated which results in recovery of some
microorganisms, and this photoreactivation represents a
significant risk to human and aquatic exposure if the reactivated
organisms are pathogenic.
As with all disinfecting agents, there is a possibility of
producing additional chemical compounds when ultraviolet
radiation is used to disinfect wastewater. The absorption of
ultraviolet radiation, especially by organic compounds, can lead
to free radical formation with resulting molecular rearrangements
and to possible mutagenesis of the microorganisms. Only
preliminary work has been done on the irradiation effects on
organic compounds; however, these initial findings indicate that
there is little chemical effect on the UV-absorbing constituents
in wastewater (Jolley, et al.; 1979). This is a marked contrast.
to chlorine and ozone, both of which produce a large number of
new organic compounds in the disinfection process (Jolley, et
a 1.; 1979).
The use of ultraviolet radiation does not produce a
residual. Consequently there is no risk to human and aquatic life
from a disinfectant residual.
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SECTION 6
NO DISINFECTION
This section investigates the alternative of discharging
wastewater that has not been disinfected. The analysis is focused
on the hazards created by discharging pathogenic organisms and
the associated hazards posed to humans since that is the primary
reason for adopting wastewater disinfection practices. This focus
on the human risks was reinforced when no discussion of hazards
to the aquatic environment from pathogenic organisms was found in
the literature.
HAZARD IDENTIFICATION
Pathogenic organisms, by definition, cause disease in human
beings. Waterborne transmission of these disease-causing
organisms can occur via four pathways:
1. direct ingestion of untreated water,
2. direct ingestion of treated drinking water,
3. ingestion of aquatic food species infected with
pathogens absorbed from contaminated waters, and
4. invasion resulting from skin contact with contaminated
water.
The first three pathways are sometimes classified as the fecal-
oral route. The second pathway described above occurs when a
drinking water treatment system fails or the integrity of the
water distribution system is violated. The fourth pathway is
likely to result in skin, mucous membrane, or urinary tract
infections but is seldom implicated in gastrointestinal illness
in the United States. Since most published research focuses on
the three fecal-oral pathways the health effects of this fourth
pathway is not well documented. Th-j risk of disease by exposure
to wastewater effluent in recreational water, especially non-
disinfected effluent, is not well established on epidemiclogical
grounds; however, recent work by Cabelli (1981) has demonstrated a
cause-effect relationship via this pathway.
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The following discussion describes the major microbes and
parasites suspected of being transferred by the fecal-oral route.
Bacteria, viruses, and protista are the most common causes of
waterborne disease outbreaks. Fungi are rarely implicated.
Bacteria.
Many of the bacterial diseases may be transmitted by
polluted water, especially those due to Salmonella and Shigella.
Salmonella typhimurium is the species most often implicated in
waterborne disease outbreaks. Symptoms include nausea, vomiting,
and fever. Death is possible without timely and adequate
treatment.
Other species implicated are Salmonel la typhi and Salmonel la
paratyphi. Salmonel la typhi is responsible for an acute disease
characterized by fever, malaise, anorexia, bradycardia (slow
heart rate),and enlargement of the spleen. Complications may
extend to the lymphoid tissues, intestinal hemorrhage, mental
dullness, and slight deafness. The fatality rate ranges from 2-3%
with antibiotic therapy. The disease is spread by food or water
contaminated with feces of a carrier or patient. A carrier is a
person who harbors the organism but remains asymptomatic.
Salmonella paratyphi may present a clinical picture similar to
Salmonel la typhi. The disease is primarily transmitted by food,
especially milk products or shellfish. The proportion of cases
that are recognized clinically is small, that is, those that
present symptoms severe enough to require medical attention.
Shigella causes a disease characterized by fever, cramps,
and abdominal pain. Shigel la sonnei is the species most commonly
isolated.
Invasive strains of enteropathogenic Eschericia coli behave
much like Shigel la. E._ col i is part of the normal intestinal
flora. Biochemically and morphologically, the enteropathogenic
strains behave the same as non-pathogenic strains. Serological
identification is necessary. Newborns are the most susceptible to
infection by enteropathogenic E_. coli, having fatalities up to
40%.
Cholera is an acute intestinal disease due to the organism
Vibrio cholerae. The disease is characterized by watery stools,
vomiting, dehydration, and circulatory collapse. In untreated
cases, the fatality rate may exceed 50%; in treated cases, the
fatality rate is less than 10% (Okun and Ponghis, 1975). Many
asymptomatic cases occur and, as mentioned previously, not all
individuals exposed to disease-causing doses of the organism
elicit disease symptoms. Primarily, the disease is transmitted by
ingestion of water or food contaminated with feces containing the
organisms.
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Another Vibrio species, Vibrio parahemo1yticus, causes
diarrhea and abdominal cramps. Though rare, the disease is most
often caused by ingestion of raw seafood.
An organism called Campylobacter fetus sub, jejuni (also
called Vibrio fetus) has been found to cause gastroenteritis.
Special culture media and modified techniques are necessary to
identify this organism so its role in waterborne disease was not
recognized until recently. It is possible that this organism was
not detected in previous gastroenteritis outbreaks because the
proper techniques and media were not used to isolate the
organism. It appears that Campy lobacter may be isolated from the
feces of patients with diarrhea as frequently as Salmonel la or
Shigel la (Sack, et al., 1980).
An organism known as Yersinia enterocolitica may cause acute
gastroenteritis, bloody diarrhea, and fever. It may also cause
pseudoappendicitis. The organism has been isolated from a wide
variety of animals including cows, beavers, and oysters. It has
also been recovered from rivers, lakes, and well water. Unlike
the previously described bacteria, this organism grows very well
at refrigeration temperatures.
Viruses.
The viruses most likely to be found in a wastewater
discharge are the human enteric virus groups Coxsackie, Polio,
Echo, Reo, Adeno, and Hepatitis A. The potential consequences of
human infection by these viruses are described below. More
detailed discussion of these consequences can be found in a
report by Benenson (1975).
Group B Coxsackie virus types 1, 2, 3, 4, and 5 have been
cultured from wastewater effluent. The Coxsackie B virus is
responsible for approximately 1/3 of the cases of~non-fatal
aseptic (nonbacteria1) meningitis. Aseptic meningitis may also be
due to Echo, adenoviruses, arboviruses, polioviruses, and the
Coxsackie A viruses which may be found in wastewater effluent.
The group B Coxsackie viruses are implicated in pleurodynia (pain
in the intercostal muscles). The disorder is characterized by
sudden onset with recurring chest or abdominal pain.
The Polio virus may cause a disease ranging in severity from
mild, to nonparalytic, to paralytic. Fever, headache,
gastrointestinal symptoms, and vague bodily discomfort are not
uncommon. Other enteroviruses, such as Coxsackie and Echo, can
produce symptoms very similar to the Polio virus. In a~rare
instance, milk has been implicated as the mode of transmission.
The p-arvovirus and reovirus-1 ike particles are believed to
be largely responsible for viral gastroenteritis. This disease is
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characterized by nausea, vomiting, diarrhea, abdominal pain, and
fever. The reovirus is frequently implicated in children's
diseases.
A large number of viruses are implicated in respiratory
disease. Death and long-term illness due to these agents is
especially prevalent in children. Adults also have a high
incidence and lose many productive hours due to these agents.
Symptoms range from cold-like chills, aches, and fever to severe
bronchitis and pneumonia. The following viruses have been
implicated: parainf luenza type 3^ adenovirus types
1,2,3,4,5,7,14,21, respiratory syncytial virus, rhinoviruses,
coronaviruses, coxsackie viruses types A and B, and echoviruses.
Presently, transfer is thought to occur by direct contact with
articles soiled by respiratory discharges, oral contact, or
droplet spread; however, waterborne transfer is possible.
The Hepatitis A virus has become a major concern in recent
years. The disease has an abrupt onset characterized by fever,
malaise (a general feeling of bodily discomfort), anorexia (loss
of appetite), nausea, abdominal discomfort, and jaundice. The
disease may follow a mild course of 1-2 weeks or last several
months. Transmission is believed to occur by way of the fecal-
oral route. Fecal contamination of recreational water then could
result in transmission of the disease.
Protista.
The most commonly implicated protozoan in waterborne disease
outbreaks is Giardia lamblia. The disease caused by this organism
is called Giardiasis. It is characterized by diarrhea, cramps,
fatigue, and weight loss. In severe cases, malabsorption may
occur. Many individuals infected with the organism are
asymptomatic. The organism exists in both a cyst and trophozoite
stage. As a cyst, it is very resistant to any type of
environmental assault, such as chlorine dosages commonly used in
water supply disinfection. Proper filtration seems to be
effective in reducing outbreaks of Giardiasis.
Another protozoan, Entamoeba histolytica, may give rise to
intestinal disease. In epidemics, it is believed that the disease
is transmitted mainly by water containing cysts from the feces of
infected individuals. Individuals who have the cysts may be
asymptomatic or have acute diarrhea with chills and fever.
Complications include the development of an amoebic granuloma
resulting in a tumor-like appearance in the wall of the large
intestine. Abscesses of the liver, lung, or brain may also
result, as well as ulceration of the skin. Like Giardia, this
organism exists in both the cyst and trophozoite stages. The
trophozoites, being very fragile, pose less threat of disease.
Filtration is effective in removing cysts.
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Naeglaria fowleri, first reported in Australia in 1965, is
associated with diseases acquired from recreational waters. It
exists in both the trophozoite and cyst stages, and its growth
may be enhanced by fecal contamination of water. The organism
causes primary amoebic meningoencephalitis, a disease
characterized by sever frontal headache, nausea, fever, vomiting,
and frequently, death. Most cases of this disease in humans have
occurred after swimming in warm, fresh water. Swimming pools,
mudholes, and mineral springs have all been implicated. Chlorine
in 10 ppm is ineffective; however, salt contents of 0.7% are
effective in controlling this organism.
Several other organisms have been implicated as being
acquired via the fecal-oral route and could theoretically be
acquired from wastewater effluent. However, their occurrence is
rare. An attempt has been- made to review the most important
organisms and to bring attention to those which will be most
frequently isolated.
SEVERITY AND FREQUENCY OF IDENTIFIED HAZARDS
The consequences from exposure to the hazards described
above have occurred many times during recorded history. Prior to
the use of disinfection and filtration in treating water supplies
those hazards and consequences were major causes of human death
and illness. But even today those consequences are occassional ly
realized. In order to provide a perspective of the potential
problem, the larger, more recent outbreaks are described in the
following discussion. The reported incidents involve the three
fecal-oral pathways described above.
Hepatitis A_^
During the period 1971-75, fourteen outbreaks of viral
hepatitis affecting 368 people occurred which were associated with
drinking water (Craun, et al., 1976). In 1975-1976, one outbreak
occurred affecting 17 individuals (Craun, et al., 1979); and it
also involved drinking water.
Salmonella typhimurium.
In 1965, in Riverside CA., more than 15,000 cases of
salmonellosis caused by Salmonella typhimurium occurred due to a
contaminated water supply that was not chlorinated. In 1978, 700-
800 people in Suffolk County, NY became ill due to Salmonella
typhimurium. W< ^tTiwater in a clogged slop sink in a catering
facility was cultured and found to be positive.
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Shigella.
In Dubuque, la., waterborne cases of Shigellosis were
documented in 1974 (Rosenberg, et al., 1976). Thirty-one of 45
cases of Shigel la sonnei were traced to swimming in water
receiving wastewater effluent from the Dubuque treatment plant;
however, state officials did not unequivocally identify the cause
of the illness as the effluent from the wastewater treatment
plant. The Dubuque wastewater was exposed to partial secondary
treatment and chlorination; and, at the time of the outbreak, the
effluent had fecal coliform counts up to 12,000,000/100 ml.
Enteropathogenic £_._ coli.
One thousand cases of diarrhea due to enteropathogenic E._
coli occured at Crater Lake National Park in 1975. The water
supply was identified as the source of the bacteria.
Vibrio cholerae.
A case of cholera occurred in Florida in 1980. The victim
had eaten approximately six dozen raw oysters in a four day
period. Two more cases of cholera followed this one, also due to
eating raw oysters. In 1978, a case of cholera occurred in
Louisiana due to ingestion of clams (Morb. Mort. Report, Dec. 19,
1980) .
Campylobacter fetus sub jejuni.
As many as 2,000 out of 10,000 town residents experienced
gastroenteritis in Vermont during a two week period in 1978
(Morb. Mort. Report, June 23, 1978). A strong association was
found between illness and the consumption of water from the town
water supply. The water was chlorinated but not filtered.
Supplementary water sources were periodically used that were not
chlorinated.
Yersinia enterocolitica.
In Europe, 10 out of 50 wells or non-chlorinated water works
have been found to contain Yersinia enterocolitica (Stern and
Pierson, 1979). In 1976, in Oneida City, NY, 220 school children
became ill due to this organism. The mode of transmission was
found to be chocolate milk. Many unnecessary appendectomies were
performed. In 1972, an elderly man became ill during a hunting
trip. The organism, Yersinia enterocolitica, was cultured from a
mountain stream from which he had drunk. Infectious levels are
not well established, but one individual in a human volunteer
study did become ill when 3.5 billion cells of Yersinia were
consumed (Stern and Pierson, 1979).
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Giardia lamblia.
Outbreaks of Giardiasis occur most commonly in the Rocky
Mountains, New England, and the Pacific Northwest. The largest
waterborne outbreak of Giardiasis occurred in New York in 1974-
75. An estimated 4,800-5,300 persons were affected. Disinfection,
but not filtration, of the water supply was practiced. A resort
town in Colorado which practices both filtration and chlorination
had a Giardia outbreak in 1979. Giardia-like illness was
described by both residents of the town and visitors to the town
(Morb. Mort. Report, March 21, 1980).
Viruses.
In July 1979, 239 cases of gastrointestinal illness were
reported among swimmers at a lake in Macomb County, Michigan.
Subsequent sampling found no violations of bacteriological
criteria for recreational waters so the etiologic agent was
assumed to be viral (Morbidity and Mortality Weekly Report,
Sept.7, 1979).
While- the reported incidents provide a qualitative measure
of the severity and frequency of the identified hazards the data
are not sufficient to support quantitative assessment. This
deficiency is even more obvious when the data are disaggregated
by the pathways described above. Furthermore, only two of the
reported incidents appear to have occurred as a result of
exposure by swimming , and only one of those waters is known to
have received a wastewater discharge. During the period 1971-
1975, the major causes of outbreaks involving municipal water
supply systems were deficiencies in the distribution system, some
involving the influx of raw wastewater. For the period 1975-1976,
the majority of outbreaks were caused by inadequately treated
water. The causative agent for 55% of 223 outbreaks since 1971
is unknown.
Additional qualitative information is provided by studies
focused on exposure via recreational waters and via occupation.
The earliest study of significance was reported in 1953
(Stevenson, 1953), and this study concluded that the incidence of
gastrointestinal symptoms was detectable among swimmers using
fresh water with total coliform densities above 2,300 - 2,400
coliforms per 100 ml. Shortly thereafter, Moore (1959) reported
no association between the incidence of polio or salmonel losis
and swimming in polluted marine waters. These studies had diverse
effects. The first study became the basis for must current
coliform standards for recreational water, ar. 1 the second study
created the impression that swimming in sewage polluted seawater
was not a health hazard. These studies have recently been
reviewed in more detail by Cabelli (1981).
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The incidence of gastrointestinal illness associated with
swimming in marine waters of varying levels of pollution has
recently been studied at three locations (Cabelli, 1981). In
general, Cabelli found:
1. the incidence of enteric symptoms were higher for
swimmers than non-swimmers,
2. only gastrointestinal symptoms increased with increasing
levels of pollution,
3. rates of illness were higher for children than for
adults,
4. symptoms declined with increased swimming time, and
5. in a body of water actually receiving raw wastewater,
residents of the area who frequently visited the beach
had much lower incidence of illness than visitors to the
area, suggesting that acclimation is a factor.
Cabelli's work is discussed further in the dose-response section
below.
Occupational exposure to wastewater was recently examined
(EPA, 1980). The purpose of this study was to determine the
effects of exposure of wastewater treatment plant workers to
viruses, bacteria, and parasites in wastewater. Experienced and
inexperienced workers and a control group not associated with
treatment plants were studied. A total of 506 individuals were
recruited for the study.
The study failed to demonstrate an increased risk to
wastewater treatment plant workers due to bacterial, viral, and
parasitic agents; however, gastrointestinal illness rates were
higher in inexperienced workers. Bacterial cultures were done
only for Salmonel la and Shigella. Several other bacterial agents
are described previously in this report which require special
techniques for isolation. By such limited selection, it is
possible that some disease causing organisms were overlooked.
Even with Cabelli's recent work, the relationship between
the concentration of indicator organisms in recreation water and
the incidence of illness is not well defined, especially in the
fresh water environment. In this study the available information
is used to illustrate a risk assessment methodology, but the
reader is cautioned to assess any results within the constraints
of the available information.
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NO DISINFECTION RISK MODEL
A no-disinfection risk model must interrelate the
probability of a consequence occurring with the quantity of
hazardous organisms discharged, the dilution effect, the dieaway
effect (or possibly, the regrowth effect), the potential insult
to the recreational water user, and the relationahip between the
magnitude of the insult and the realization of the consequences
(the dose-response relationship). Each of these subjects is
discussed below in the sequence given above, and the discussion
ends with a synthesized model.
Magnitude of Discharge Hazards—
Conventional wastewater treatment provides reductions in
microbes by way of the following processes: sedimentation,
aeration and sedimentation, and natural dilution and die-off
during surface water discharge. The typical levels of
microorganisms entering a treatment plant are given in Table 16.
TABLE 16. INFLUENT CONCENTRATION RANGES FOR
PATHOGENIC AND INDICATOR ORGANISMS (EPA,1979a)
Organism Number/lOOml
minimum maximum
Total Coliforms 1,000,000 46,000,000
Fecal Coliforms 340,000 49,000,000*
Fecal Streptococci 64,000 4,500,000
Virus 0.5 10,000
* - Apparently, the samples containing the maximum fecal
coliform levels were not analyzed for total coliforms
because fecal coliform levels can never exceed the total
coliform levels.
Table 17 shows the percent reduction of microorganisms by
primary and secondary treatment. Reductions in bacteria and
viruses up to 99% by primary and secondary treatment are cited
in a recent EPA publication (EPA, 1979b). Presently, the
quantitative measurement of fecal coliforms serves as an index to
the presence of fecal contamination. Treatment of water, it is
assumed, removes pathogens in proportion to the reduction in
indicator organisms.
A relationship has been reported (Kerr and Butterfield,1943)
between coliforms and typhoid bacteria in wastewater. The results
were approximately 27.5 Salmonel la sp. (Salmonel la other than
typhi) per 100,000 coliforms, and considerably fewer Salmonel la
typhi per 100,000 coliforms.
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TABLE 17. MICROORGANISM REDUCTIONS
BY CONVENTIONAL TREATMENT PROCESSES
(Okun and Panghis, 1975; Craun, et al., 1976)
Microorganism
Primary treatment Secondary treatment
remova1,% remova1,%
Total coliforms <10
Fecal coliforms 35
Shigella sp. 15
Salmonella sp. 15
Escherichia coli 15
Virus <10
Entamoeba histolytica 10-50
90-99
90-99
91-99
96-99
90-99
76-99
10
Using this relationship and assuming median Table 17
reductions in microbes, the levels of microbes shown in Table 18
can be expected after secondary treatment.
TABLE 18. SECONDARY EFFLUENT RANGES FOR
PATHOGENIC AND INDICATOR ORGANISMS
Organism
number/lOOml
minimum
maximum
Total Coliforms
Fecal Coliforms
Fecal Streptococci*
Viruses
Salmonella sp.
45,000
11,000
2,000
0.05
12
2,020,000
1,590,000
146,000
1,100
570
* assuming removal efficiencies for fecal streptococci
similar to the fecal coliform removal efficiencies.
Dilution Effect—
Dilution of a wastewater discharge will reduce the
concentration of the hazardous organisms. This dilution effect
can be incorporated into the model by multiplying the quantity of
discharged organisms by a dilution factor which is defined for
streams as the ratio of the discharge flow to the sum of the
discharge and receiving stream flows. For lakes receiving
discharge the dilution factor would be the ratio of the discharge
flow to the lake outflow. This dilution factor will calculate an
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organism concentration for the lake that will then be diminished
by the dieaway factor discussed below. The equilibrium
concentration in the lake will then be a result of both dilution
and dieaway.
Die-away of Pathogen Discharges—
The natural processes of die-off and inactivation will
eventually reduce pathogen levels substantially. A bacterial die-
off study has been reported (Dutka, et al., 1980) in Canada. A
200 ml aliquot of a four-day old Serratia culture was inoculated
into a stream. The suspension contained approximately 100 million
cells/ml. Serratia is a bacterium closely related to Klebsiel la,
a common isolate in normal human stools. The organism was
recovered for a period of up to 22 days and traced for a distance
of 20 kilometers (approximately 12.4 miles). In addition to the
river studies, studies were done in lakes, and survival times up
to 28 days were found for E^ coli, enterococcus, and Salmonel la
thompson. When pollution levels in lakes were compared, it was
noted that faster die-off times were observed in less polluted
lakes. This probably occurs because there are less nutrients and
organic carbon available in the less polluted bodies of water.
The bacteria, therefore, are not able to reproduce at a rapid
rate. Any inferences from a study of this type must consider the
variations in numbers of microorganisms between treated
wastewater effluent and a broth culture as utilized in the
study. There are many competitive organisms in wastewater
effluent, and nutrients vary from those supplied in a laboratory
broth, depending on the nature of the discharge.
In arid and semi-arid areas, such as Denver, wastewater
effluent may make up a large portion of stream flow. For example,
in Clear, Sand, and Cherry Creeks in the Denver area, flow from
wastewater may constitute 40-70% of stream flow (EPA, 1977).
Dilution effect's in this case are minimal, and die-away will be
the primary mechanism reducing pathogen populations.
Numerous mathematical models have been proposed for
modelling bacterial die-away; however, models based on first
order kinetics are most frequently selected for application. The
first order model for die-away in streams is:
N = NQ e(~kt)
where NQ = the initial concentration of microbes discharged
into the stream, and
N = the concentration of the microbes t time units
after discharge into the stream.
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A first order model for die-away in standing water bodies (i.e.
lakes) is:
N = NQ/(l+ktd)
where NQ = the concentration in microbes in the water body's
inflow,
N = the concentration of microbes in the water body's
discharge, and
td = the hydraulic detention time in the water body
based on the water body's discharge.
The rate constant k can be determined from a dieaway study
of typical lakes or streams in a planning area. For example, if a
summer dieaway stream study found a 30% reduction of the
microorganisms in two days, k would be:
k = (1/t) In (No/N)
k = (1/2) In (100/70)
k = 0.18 per day.
Rate constants may also be found in the literature
(Bitton,1978 and Berg, 1978).
Potential Insult—
The phrase "potential insult" refers to the potential for a
human to ingest pathogenic organisms. The method of expressing
this factor will depend on the type of dose-response model used.
For example, if the dose-response model is defined as the number
of organisms required to elicit a response in a given fraction of
insults then the potential insult must be stated in terms of
total organisms ingested. This quantity can be calculated by
multiplying the organism concentration by the amount of water
ingested while swimming which in this study was assumed to fall
in the 50 - 500 ml range.
On the other hand, a dose-response model may relate
probability of illness to organism concentration. In that case,
the potential insult is incorporated into the dose-response model
so this element does not have to be estimated separately.
Dose-Response—
Data relating dosage levels of microbes to risk of disease
has been available for only the last few years. Available data
seem to indicate that a considerable number of organisms are
required to elicit a response for several pathogens. For example,
the most severe of all diarrheal diseases, cholera, is reported
to require 100,000,000 organisms to elicit a response, and of
83
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those challenged, only 1-10% will develop clinical
manifestations (Sack, et al., 1980). On the other hand, it has
also been reported that 10-200 Shigella can elicit a response in
those challenged (Sack, et al., 1980). More data collection needs
to be done on Shigel la since this is the bacterial pathogen which
apparently requires the lowest number of ingested organisms to
initiate a disease process.
The infective dose for pathogenic protozoa is reported to be
small in some cases. For example, one Giardia cyst has been
reported to be infective. However, in an experiment where adult
humans were challenged with ten organisms, 76-100% did not become
ill (EPA, 1979a). The infective dose for Entamoeba histolytica is
likewise not well established.
Knowledge concerning the infective dose of viruses is
perhaps the most poorly established. Many virologists believe
that one active virus is sufficient to initiate a response
(Mahdy, 1979).
Two dose-response models were found in the literature for
Salmonella (Mechalas, et al., 1972; EPA, 3979b)- Both models are
represented graphically in Figure 4, and the lack of agreement
between the two models is substantial and obvious. For example,
10 Sa1mone1 la sp. organisms per liter causes a risk of 1
illness in 5 billion exposures (EPA, 1979b), or a risk of 1
illness in 3300 exposures (Mechalas,1972). However, the levels of
Salmone1 la in secondary treatment effluent, as shown in Table 18,
would seem to pose a low risk of illness.
In order to use a dose-response model with water quality
data as an input, the dose units should be indicator organisms
instead of pathogens. The Mechalas (1972) models include coliform
and fecal coliform curves in addition to the Sal mone 1 la and virus
curves. The implication in those models that the ratios between
the indicator species and the pathogens are constant is debatable
and raises doubt about the models' accuracy.
Because of the two problems cited above with the reported
Salmone1 la models, those models were not used in the no
disinfection risk model.
Cabelli (1981) has also presented quantitative regression
models relating the occurrence of gastrointestinal symptoms per
1000 swimmers to the concentration of an indicator organism. His
work also evaluated several indicator organisms, and he found far
better correlation using enterococcus or £_._ coli than using the
more commonly used coliform and fecal coliform indicator
organisms. In fact, the correlation for the coliform indicators
are so poor that a useful symptom-col iform model cannot be
synthesized. The dependence of the recommended Cabelli model on
Enterococcus input-data greatly diminishes the model's usefulness
84
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since very little Enterococcus data are available. Cabelli
presents several regression models, and the model best suited to
this study's approach is:
y = 11.85 log x - 0.58
where y = the number of swimmers developing gastrointestinal
symptoms per 1000 swimmers and
x = the mean enterococcus density per 100 ml.
Assembled Model—
The relationships presented above can be assembled into the
following mathematical models. Using Cabelli's dose-response
model and assuming the wastewater is discharged to a stream the
model becomes:
y = 11.85 log[A Qe exp(-kt)/(Qe + Qs)] - 0.58
where y = the number of gastrointestinal symptoms expected
per 1000 swimmers,
A = the enterococcus concentration in the wastewater
discharge in organisms/100 ml,
Qe = wastewater flow rate in volume per unit time,
Qs = receiving stream flow rate in the same units as Qe,
t = time of travel in the stream from the point of
discharge to the point of exposure in time units, and
k = the rate constant as described above with the same
units as t to the minus 1 power.
The model for discharge to a standing water body would be:
y = 11.85 log[A Qe/(Qd(l + ktd))] - 0.58
where Qd = standing body discharge flow rate in the same
units as Qe,
td = hydraulic residence time in the water body, and
y, A, Qe, and k are defined above.
Probability factors can be added to the model by estimating
probabilities for the input variables and then integrating them
with the Cabelli model's statistics. The use of the model shown
above is illustrated by an example in Section 8.
86
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SECTION 8
ENERGY AND COST CONSIDERATIONS
ENERGY
Chlorination
The on-site energy requirements for the Chlorination of
wastewater are small compared with those of ozonation or
ultraviolet irradiation. An analysis of energy requirements for
Chlorination of municipal water supplies has recently been
published by Clark (1981). In this analysis of Chlorination
energy demands, values for the energy efficiency, in kWh/kg Cl2/
should vary little between wastewater and drinking water plants
of the same size. Hence, the energy requirements listed in
Clark's paper will be used as a good approximation for the energy
efficiency of the Chlorination process in wastewater treatment
plants. Clark reports that the energy efficiency is a function of
the disinfectant capacity of the plant. For example, the energy
efficiency is given as 1.88 kWh/kg C12 at a design capacity of
37.9 kg CL.2/d, and a value of 1.41 kWh/kg Cl2for a design
capacity ot 75.7 kg Cl2/d. Doses in the range 4 - 16 mg Cln/L are
typical of wastewater treatment plants. For a dose of 6 mg/L,
assuming 100% C12 absorbed, a 3790 m3/d plant would require 22.5
kg Cl2/d, i.e.,(5790 m3/d)x (6mg/L)x(kg/106 mg)x (1031/m3) = 22.5
kg Cl2/d. Based upon the analysis of Clark, (Table 10, Clark,
1981), a plant with a capacity of 22.5 kg Cl2/d would have an
energy efficiency of about 3.7 kWh/kg C12.
Only on-site energy requirements are included in this study.
Off-site power consumption associated with chlorine manufacture
and the energy involved in transport are not included in energy
calculations. This approach results in an apparent lower energy
cost for chlorine as compared with alternative disinfectants
that must be produced on-site.
Ozonation
In this section the energy requirements for wastewater
disinfection by ozonation are discussed. Energy-intensive steps
involved in ozonation are discussed first, followed by some
87
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comments on how the quality of the influent may affect the energy
requirements for ozonation of water. Energy data for ozonation of
wastewater and municipal water supplies will be presented and
compared, in the context of differences in water quality, as well
as the variance associated with different equipment. This
analysis will be used to provide a range of energy values, in
kWh/kg 03, for ozonation of wastewater. Finally, an estimate of
required kWh per unit of treated wastewater will be obtained for
a typical range of values for the mass transfer efficiency.
Energy-Intensive Steps in the Ozonation Process—
Though air is by far the most common feed gas for ozonators
treating municipal water supplies, in wastewater treatment plants
in the U.S (those either operating or in various stages of
development) enriched oxygen (air enriched by oxygen) is roughly
as common a feed gas as air. In wastewater treatment, where
oxygen may be generated on-site for oxygen-activated sludge, the
ready availability of oxygen as a feed gas offers the possibility
of higher levels of ozone production for the same energy input to
the ozonating equipment. However, if the feed gas is air or
partially-enriched oxygen (associated with an ozonation process
in which oxygen ^s recycled), air pretreatment, involving
compression, refrigeration, and drying, is essential (Rosen,
1976).
Manufacturers of ozone generators may sometimes promote
values for energy requirements obtained under the most ideal
conditions. However, the energy efficiency of a specific ozone
generator is a function of many variables, including: power
level, condition of the generator, cooling fluid temperature, as
well as the flow rate, moisture content, and oxygen content of
the feed gas. Of. course, the energy efficiency wi 1 1 also vary for
different model generators. The effect of these variables on
energy efficiency of ozone generation has been discussed in
detail by Carlins (1981).
In addition to the energy cost of feed gas pretreatment and
ozone generation, the overall energy efficiency will depend
largely on the mass transfer efficiency of the contactor, as well
as the dosage level required for adequate disinfection. Finally,
additional energy is usually expended to operate an ozone
destruct unit which prevents the contactor off-gases from raising
ambient ozone levels beyond acceptable levels.
Qualitative Comparison of Energy Requirements for Ozonation of
Wastewater and Potable Water—
The rather limited energy data available for wastewater
treatment plants using ozone suggest that the much larger body of
published data for the treatment of municipal water supplies
should also be considered. The energy efficiencies, in kWh/kg O3/
should be very similar; however, some cautionary observations are
appropriate. First, the dosage levels required for adequate
88
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disinfection are higher in wastewater treatment, not only because
of the presence of higher levels of microorganisms but also the
presence of higher levels of ozone demand. The presence of
various organic and inorganic reactants has led to a second
problem. The mass transfer efficiency at some of the early U.S.
wastewater treatment plants has turned out to be considerably
lower than expected. This is likely due, in part, to the then
embryonic state of development of the ozonation process in
wastewater treatment. The use of PVC piping and inappropriate
epoxies has been associated with serious leaking in some
contactors (Jain, et al., 1979; Rakness and Hegg, 1980) which
would cause low transfer rates. Moreover, there is some evidence
that the mass transfer capabilities of contactors are affected by
the level of reactant pollutants (Rakness and Hegg, 1980). Thus,
care must be taken that in a laboratory pilot study of a
particular diffuser design, the water quality be as close as
possible to that at the proposed plant. A low mass transfer
efficiency, associated with an under designed or poorly designed
contactor, implies a higher value of kWh/unit treated water, as
well as higher levels of ozone in the contactor off-gases which
enter the destruct unit. Finally, ozonation systems used in the
treatment of municipal water supplies use air as a feed gas
almost exclusively, so the data from such systems will have to be
adjusted appropriately for comparison to wastewater plants which
often use oxygen as a feed gas.
Energy Efficiency (kWh/kg 03) for Ozonation of Wastewater and
Potable Water—
Air as Feed Gas.—Table 19 presents the energy efficiency,
in kWh/kg 03, for twelve European and Canadian plants treating
potable water. The data were published in a 1978 EPA report
(Miller, et al., 1978) which included the results of a
questionnaire mailed to numerous municipal treatment plants
throughout Europe and Canada. In Table 19 only the data from
those plants which provided a breakdown of energy consumption for
various stages of the ozonation process are presented. The
average values for air preparation, generation, contacting, and
ozone destruct are, respectively, 7.8, 19.3, 6.6, and 8.7 kWh/kg
63. The average total energy efficiency is 29.5 kWh/kg 03, where
some of the plants did not report energy consumption values for
contacting or ozone destruct. Note that while the absence of an
additional energy requirement for the contactor is consistent
with a system driven by the positive pressure of the off-gases
from the generator, a destruct unit would require additional
power. An examination of the data in Table 19 reveals no clear
pattern associated with an economy of scale; of course, various
"noise" factors discussed above may effectively mask such a
pattern.
Comparable data for wastewater treatment plants is quite
sparse. Many U.S. plants have had severe difficulties getting up
to full operational capability, with shake-down periods commonly
89
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TABLE 19. ENERGY UTILIZATION FOR EUROPEAN AND CANADIAN PLANTS
TREATING MUNICIPAL WATER SUPPLIES (Miller, et al.,1978);
AND ONE U.S. PLANT TREATING MUNICIPAL WASTEWATER
(Rakness and Hegg, 1980)
Plant Capacity
Air
Prep.
Ozone
Gen.
kg 03/d
Lock Turret
Durleigh
Cantineweg
Linz
Biel
Toulon
Villeneuve-
la Garonne
Flins
Drummondville
lie Perrot
Pierref ords
Sherbrooke
AVERAGES
Estes Park, CO
168
192
120
27
130
36
96
264
41
9
250
163
25.3
2
12
3
14
2.3
11
12
12
2.2
9
6
7.9
7.8
17.3
22.6
18
21
21
17
26
18
18
15.4
21
12
21.2
19.3
17.3
Ozone Ozone
Contact Destruct
kWh/kg 03
6.6 5.3
6
0.6
10
3.7
12
4.4
15
6.6 8.7
0 4.3
Total
37
36
24.
45
23
47
30
30
22
45
18
29.
29.
38.
5
1
5
9
extending several months or a few years beyond the scheduled
start-up date. These delays have been due to ozone-related
problems, as well as other factors. However, significant energy
data have been published for the Upper Thompson Sanitation
District (UTSD) wastewater treatment plant in Estes Park,
Colorado. Data for the UTSD plant at 70% of actual capacity,
presented in Table 19, are in good general agreement with the
energy efficiency of the European plants. The designers of the
UTSD plant have noted a significant difference between the
manufacturer's specifications for energy efficiency and actual
operational values (Rakness and Hegg, 1980). They attribute this
discrepancy to a faulty dew point measuring device which caused
the moisture of the generator feed air to exceed the
manufacturer's minimum specified value. They also note that if
the UTSD ozone generator had achieved rated energy efficiency,
the total energy efficiency would have been about 34 kWh/kg O3 at
70% capacity, instead of the observed value of 38.9 kWh/kg O3-
Their projected value brings the UTSD data into even closer
agreement with the average total energy efficiency of 29.5 kWh/kg
03 for the European and Canadian plants.
90
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p_2 as Feed Gas.—As discussed above, there are very limited
energy—utilization data for wastewater treatment plants using air
as feed gas for ozonation; however, even less data are available
for systems using oxygen as the feed gas. The Southwest
Wastewater Treatment Plant in Springfield, Missouri, has been
operating successfully for about three years, with careful
monitoring of costs (Letterman,1981) and is the only source of
energy data for oxygen-fed ozonation found in the literature.
Oxygen is generated on-site for an oxygen activated sludge
system, as well as for ozone generation. The Springfield system
uses a once-through oxygen process, in which the feed gas is 98%
pure oxygen, the generators provide a 3% conversion of oxygen to
ozone, and off-gases are channeled directly to the biological
system. There is no ozone destruct unit. Ozone mass transfer
efficiency has typically been in the range of 80 -90%. Table 20
summarizes the energy requirements of the Springfield plant. The
energy requirement for oxygen generation, 11 kWh/kg 03, has been
obtained from the reported values of .33 kWh/kg 02 produced
(Breitback,1981) and the 3% conversion efficiency for ozone
generation from oxygen. The energy efficiency of the Springfield
ozone generators, 10.9 kWh/kg 0-,, is consistent with the energy
efficiency for air-fed systems (Table 19) and the approximate
two-fold increase in efficiency associated with oxygen-fed
systems (Rice, 1980).
TABLE 20. ENERGY UTILIZATION FOR AN OXYGEN FED
WASTEWATER OZONATION PLANT (Letterman, 1981)
Plant
Capacity 02 Generation Blowers Generators Total
& Mixers
Springfield
MO
kg 03/d
1360 11
kWh/kg 03
0.53 10.9
22.4
Energy Utilization in kWh/Unit treated water—
The energy utilization, in kWh/unit treated water, will be a
function of energy efficiency (kWh/kg 03), the mass transfer
properties of the system for ozone, and the water quality. In
quantitative terms,
Energy utilization(kWh/L) = energy efficiency(kWh/kg 03) x
dosage(kg 03/L) /
mass transfer efficiency
Doses in the range of 4 - 8 mg 03/L are typical of plants which
do not have significant amounts of industrial wastes in their
91
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influent waters (Rice, 1980). Table 21 provides energy
utilization values, in kWh/L treated water, corresponding to an
energy efficiency of 22, 30 or 35 kWh/kg 03, a mass transfer
efficiency of 50% or 85%, and an ozone dose of 1 mg/L.
TABLE 21. ENERGY UTILIZATION PER UNIT OF OZONATED WASTEWATER
kWh/1 treated water/(mg 03 absorbed/1)
Mass Transfer Efficiency
Energy Efficiency
kWh/kg 03 50% 85%
22
30
35
4.4x10 i 2.6x10 ;?
6.0xlO~3 3.5x10 j?
7.0xlO~5 4.1xlO~b
Use of Table 21 is straight forward. For example, at 30 kWh/kg
03, 85% mass transfer efficiency, and an absorbed ozone dose of 6
mg/L, a 3,785 m3/d plant would require ((3,785,000 i/d) x
(3.5xlO~5 kWh/l/(mg/L)) x 6 mg/L =) 795 kWh/day.
Ultraviolet Radiation
Since the early 1900's ultraviolet light has been used, on
occasion, to disinfect small quantities of drinking water
(Scheible, 1980). The lack of residual disinfectant in the
treated water, though generally viewed as a disadvantage for the
treatment of potable water, may be an advantage for wastewater
treatment. Various small pilot studies have explored the
practicality of using UV for the treatment of wastewater. The
first full-scale ultraviolet disinfection field study was
conducted during a 13-month period in 1978-1979 at the Northwest
Bergen County Water Pollution Control Plant in Waldwick, New
Jersey (Scheible, 1980). The UV apparatus at the plant was
capable of treating the full plant flow of 15,000-30,000 m3/d
(3.9-7.9 mgd). Plant data were used to estimate energy
requirements for typical wastewater treatment plants. Energy
requirements of 4.4 x 104 kWh/yr, 4.4 x 105 kWh/yr, and 4.4 x
10b kWh/yr were estimated for plants with design capacities of
1 mgd, 10 mgd, and 100 mgd, respectively. These values imply an
energy utilization of 3.2 x 10"^ kWh/L of treated water. These
values for the energy requirements and energy utilization are
based upon doses required to mtat current effluent standards.
Existing standards do not include consideration of repair
phenomena which may occur in bacteria subsequent to disinfection.
It has been shown that the repair of bacterial DNA damaged by UV
irradiation is enhanced if the organisms are subsequently exposed
to visible light. The increase in dose required to compensate for
the photoreactivation reaction would depend upon the season and
92
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geographical location of the plant. A factor of 2x increase in
dose was used by Scheible(1980) as an approximation of the dose
needed to compensate for photoreactivation which would increase
the energy utilization to 6.4 x 10~5 kWh/1 of treated water.
Comparison of On-Site Energy Requirements
An example comparison of the on-site energy requirements for
the alternative disinfection processes considered can be
developed using a 3790 m3/d wastewater treatment plant. For
wastewater treatment, absorbed doses in the range of 4-8 mg 03/L
or 4-16 mg C^/l are typical of plants which do not have
significant amounts of industrial wastes in their influent
waters. For the example comparison, absorbed doses of 6 mg 03/L
and 10 mg C12/1 are assumed to be equivalent for the purposes of
disinfection. An ozone transfer efficiency of 85% is also
assumed. As discussed above, the energy efficiency for air-fed
ozonation is about 30-35 kWh/kg 03 for a wide range of plant
sizes. For the sake of comparison, a value of 35 kWh/kg 03,
suggested by the data obtained from the 1.3 mgd UTSD plant in
Estes Park, Colorado will be assumed. The energy utilization of
oxygen-fed ozonation plants is assumed to be 22 kWh/kg 03. The
energy requirements, based on the above assumptions, for a 3790
m3/d plant are shown in Table 22.
TABLE 22. EXAMPLE COMPARISON OF ENERGY REQUIREMENTS
FOR ALTERNATIVE DISINFECTANTS
kWh/d
Disinfection Alternative Energy Requirement
kWh/d
Chlorination 71
Ozonation
Air-fed 932
Oxygen-fed 586
Ultraviolet
w/o photoreactivation 121
w/ photoreactivation 242
As discussed above, the energy requirement for Chlorination
will vary with treatment plant size. For smaller plants the
requirement could be substantially larger than shown in Table 22,
and for larger plants the requirement will decline slightly from
the value shown.
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COSTS
This section discusses capital and 0 & M costs for
disinfection by chlorination and ozonation and reduces them to
costs per 1,000 cubic meters. Sensitivity to some of the
variables is estimated.
Chlorine
Capital Costs for Chlorine Facilities—
The 1980 EPA study of wastewater treatment construction
costs identified 92 construction projects that included
chlorination for disinfection. After adjusting all prices to the
fourth quarter 1978, cost ($ million) was regressed on design
flow (in mgd) to produce the following estimating equation for
cost as a function of quantity (EPA FRD-11, 1980):
Cost = (6.33 x 104)
x (Q°«65)
In order to bring this equation up to date, inflation to the
first quarter of 1981 must be added. The EPA small city
conventional treatment plant coat index for 1 mgd and 10 mgd
chlorination systems and the large city advanced treatment index
for the 100 mgd plant were used to add the inflation. The
relevant indices are shown in Table 23 (EPA, 1981).
TABLE 23. CONSTRUCTION COST INDICES
PLANT
SIZE
m^/day
3790
37900
379000
1978
INDEX
152.0
152.0
157.0
1931
INDEX
174.7
174.7
191.2
PERCENT
CHANGE
%
15.0
15.0
21.8
The wide variations in percent increase do not actually
reflect differing inflation rates, but rather, a peculiarity in
the base years. The 1981 figures are national averages while the
1978 figures are based only on data from Kansas City and St.
Joseph, MO. The Kansas City data were within 1% of the national
average in 1978, but the St. Joseph data was 5% above the
average.
Non-construction project costs (design costs) also need to
be added to costs. Administration, legal costs, architects and
engineering fees, inspections, and contingencies add another 28%
94
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to new projects or 36% on enlargement or upgrading projects (EPA,
FRD-11, 1980). A median figure of 32% for non-construction costs
is used in this analysis.
Capital costs shown in Table 24 for a chlorine disinfection
system are the product of the basic estimate, inflation, and the
non-construction cost ratio.
TABLE 24. CHLORINATION CAPITAL COSTS
PLANT
SIZE
m^/day
3790
37900
379000
1978
COSTS
$
63,000
294,000
1,367,000
INFLATION
COSTS
$
9,450
44,100
298,000
DESIGN
COSTS
$
23,550
107,900
533,000
1981
COSTS
$
96,000
446,000
2,198,000
The peculiarities of individual cases lead to large
variations around these averages. The EPA data on individual
construction jobs show variations by more than a factor of three
both up and down in about 20% of their cases. Area costs of
construction vary from the average by up to 20% even if the
highest and lowest cost cities (New York and Charlotte, NC) are
excluded.
Annual costs for several interest rates with twenty year
amortization periods are shown in Table 25 for three sizes of
treatment plants. In order to estimate annual costs it is
necessary'to select an amortization period, an amortization
TABLE 25. ANNUAL CHLORINATION CAPITAL COSTS
INTEREST
RATE
%
0
0
3
6
9
12
15
AMORTIZATION
PERIOD
yrs
1
20
20
20
20
20
20
3790
$96,000
4,800
6,432
8,352
10,464
12,864
15,360
PLANT SIZE
m3/d
37900
$446,000
22,300
29,882
38,802
48,614
59.764
71,360
379000
$2,198,000
109,000
147,226
191,226
239,582
294,532
351,680
95
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interest rate, and a utilization rate. The latter factor
increases the cost per unit of wastewater treated because of
unused capacity.
A twenty year amortization period is assumed for this
analysis. The interest rate on good quality municipal bonds was
approximately 11% in mid 1981. The rate is held down by the
federal and state government tax exemption, and it also fails to
fully reflect increasing risks as communities approach the limits
to their bonding capacities. On the other hand, these rates are
pushed abnormally high by the fear of inflation and the federal
effort to end inflation. Nine percent (9%) was selected as a
reasonable interest rate for amortizing a wastewater investment
in this analysis. Other interest rates might vary the
amortization per year by up to a factor of two.
Unit capital costs for chlorination are shown in Table 26
for three utilization rates. The annual amortization costs for 9%
interest and a 20 year amortization period were used as bases for
the unit costs shown. Cost variations from interest rates and
regional differences in construction costs would scale this whole
table upward or downward proportionately. For example, a six
percent interest rate would reduce all capital costs by 20
percent.
TABLE 26. UNIT CHLORINATION CAPITAL COSTS
PLANT
SIZE
m3/day
3790
37900
379000
ANNUAL UTILIZATION
COST RATE
$ %
* 10,464 100
80
60
48,614 100
80
60
239,582 100
80
60
ANNUAL
COST
$/1000
7.56
9.46
12.61
3.51
4.39
5.86
1.73
2.16
2.89
UNIT
m3
Operating and Maintenance (O&M) Costs Using Chlorine—
The C & M costs for disinfection by chlorine can be
categorized as chemical, labor, supplies, and power. Chlorine
alone will typically account for half or more of all O & M costs.
96
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Chlorine dose can vary with the wastewater being treated,
and a common figure in the literature appears to be 10 mg/L
(Opatken, 1978). The 1980 WPCF survey suggests a median dose of
just over 6 mg/L (WPCF, 1980). The costs per unit of treated
wastewater for these two dosages are shown in Table 27.
TABLE 27. CHLORINATION DISINFECTANT COSTS
PLANT
SIZE
m3/day
3790
37900
379000
PRICE
$/kg
0.24
0.24
0.18
DOSE
mg/L
6
10
6
10
6
10
ANNUAL
USAGE
kg/yr
8,300
13,800
83,000
138,000
830,000
1,380,000
UNIT
COST
$/l,000 m3
1.44
2.40
1.44
2.40
1.08
1.80
Price for chlorine also varies. In the Denver area, railroad
tank car lots can be bought for $0.18/kg (Metropolitan Denver
Sewage District No.l) and 908 kg cylinders for about $0.24/kg
(Denver Water Dept.). Smaller (68 kg) cylinders appear to cost
more than twice as much and are probably not cost effective even
for a 3790 m3/d plant. Median price for chlorine in the 1980 WPCF
survey was $0.26/kg (WPCF, 1980). The unit costs shown in Table
27 are based on the assumption that the smaller plants will use
908 kg cylinders and the large plant will purchase chlorine in
railroad tank cars.
Labor is the second category of operating cost. Maintenance,
labor and supervision are assumed to be proportional to plant
investment totaling 1% of cost. Labor cost is assumed to be $10
per hour. Labor costs per unit of wastewater treated are shown in
Table 28.
TABLE 28. CHLORINATION LABOR COSTS
PLANT
SIZE
m /day
3790
37900
379000
ANNUAL
LABOR
hrs
460
1,840
7,360
ANNUAL
COSTS
$
4,600
18,400
73,600
UNIT
COSTS
$/l,000
3.33
1.33
0.53
m3
97
-------
Power costs are primarily for heating, lighting and
ventilation. While electricity can still be bought at 4-5 cents
per KWh, new capacity generally costs at least 6 cents and in
some cases more. Power costs per unit of wastewater treated are
shown in Table 29.
TABLE 29. CHLORINATION POWER COSTS
PLANT
SIZE
m3/day
3790
37900
379000
ANNUAL
CONSUMPTION
kWh
10,000
20,000
20,000
ANNUAL
COSTS
$
600
1,200
1,200
UNIT
COSTS
$/l,000 m3
0.43
0.09
0.01
Supplies for maintenance are estimated at 1% of capital
cost?
Total 0 & M cost using chlorine for disinfection are the sum
of the components. At 3790 m3/d, these costs are most sensitive
to labor costs. Sensitivity to the cost of chlorine is dominant
in the larger plants. Chlorination operating and maintenance
costs are summarized in Table 30.
TA^LE 30. CHLORINATION 0 & M COSTS SUMMARY
$/l,000 m3
Disinfectant
Labor
Power
Supplies
3790
1.44 - 2.
3.
0.
0.
40
33
43
08
PLANT SIZE
m3/day
37900
1.44 - 2.40
1.33
0.09
0.04
379000
1.08 - i.
0.
0.
0.
80
53
01
02
Totals
5.28 - 6.24 2.90 - 3.86 1.64 - 2.36
As with capital costs, 0 & M costs per unit actual ly
disinfected can be expected to increase with underuti1ization.
The effects of underuti1ization on Chlorination O & M costs is
shown in Table 31 with the labor costs per unit of wastewater
treated assumed to be constant.
98
-------
TABLE 31. UTILIZATION EFFECTS ON CHLORINATION 0 & M COSTS
$/l,000 m3
PLANT
SIZE
m3/day
3790
37900
379000
PERCENT UTILIZATION
60 80
6.58 - 8.18 5.77 - 6.97
3.95 - 5.55 3.29 - 4.49
2.38 - 3.58 1.92 - 2.82
100
5.28 - 6.24
2.90 - 3.86
1.64 - 2.35
Chlorination Costs Summary—
Total chlorine disinfection costs are summarized in Table 32.
The chlorination costs shown are fairly consistent with similar
costs cited in the literature. Van Note (1978), Opatken (1978),
and Gupta (1976) all show higher 0 & M costs after adjusting to
1981 prices because they used higher chlorine prices. Opatken
also assumes a higher labor input than used in this analysis.
Opatken and Gupta estimate costs for 1.3 mgd and 2.0 mgd plants,
respectively, and apparently used the more expensive 68 kg
cylinders. The estimates of capital cost amortization by Van
Note, Opatken and Gupta are lower than the estimate in this
analysis because they used a lower interest rates. Gupta's lower
capital costs are offset by his inclusion of additional capital
for dechlorination and post-aeration. The differences reflect
primarily the differences in conditions when the studies were
done. A fourth study by Nail (1980) did not break down the costs,
but found ^total chlorine disinfection costs similar to the
results in this analysis for the small plant and somewhat lower
costs for the two larger plants.
TABLE 32. CHLORINATION COSTS SUMMARY
$/l,000 m3
PLANT
SIZE
m3/day
3790
37900
379000
PERCENT UTILI
19.19
9.81
5.27
60
-
20.79
11.41
6.47
80
15.23
7.68 -
4.08 -
ZATION
16.
8.
4.
43
88
98
12.84
6.41
3.37
100
- 13.
7.
4.
80
37
08
Notes
- minimum unit cost corresponds
- maximum unit cost corresponds
- interest rate = 9%
- amortization period = 20 years
to
to
6
10
mg/L Cln
mg/L Cl.
dosage
2 dosage
99
-------
Ozone
Capital Costs for Ozone Facilities—
Ozone cost experience in this country is extremely limited;
however, in Europe, there is substantial experience with drinking
water treatment rather than wastewater. As a result, cost
estimates for ozone treatment of wastewater have not been
confirmed by substantial experience and are subject to a wider
band of uncertainty than chlorine disinfection costs.
The major controlling factor for ozone disinfection cost is
the amount of ozone that must be applied to the wastewater. This
in turn is the product of two factors, the amount needed for
disinfection and efficiency with which an applied dose becomes
available.
The amount of ozone required for disinfection will, of
course, vary with the wastewater and the standards for
disinfection. Venosa, et al.(1978) found that ozone doses of 4
mg/L of wastewater were sufficient to meet EPA standards of
disinfection with reasonably good secondary effluent, and 5 mg/L
of water provided sufficient safety margin.
The Venosa experiments found absorption rates between 50 and
90 percent for ozone (Opatken, 1978). Operators of ozone
disinfection plants report absorption rates as low as 40 percent.
Ozone generation requirements for a 5 mg/L dose at varying
absorption rates are shown in Table 33.
TABLE 33. OZONE PRODUCTION REQUIRED
TO PRODUCE AN EFFECTIVE DOSE OF 5 MG/L
kg/day
PERCENT
ABSORBED
%
90
70
50
40
APPLIED
DOSE
mg/L
5.55
7.14
10.00
12.50
3790
21
27
38
47
PLANT SIZE, mj/d
37900
209
268
377
472
379000
2092
2679
3768
4722
In 1978, EPA published construction costs for ozone
generation and contactor systems using an air feed for the
smallest systems and pure oxygen for larger systems (Gumerman,
1978). Ozonation capital costs with design and inflation costs
included are shown in Table 34. The minimum cost shown
corresponds to a contactor efficiency of 90%, and the maximum
cost corresponds to a contactor efficiency of 40%.
100
-------
TABLE 34. OZONATION CAPITAL COSTS
PLANT
SIZE
m3/d
3790
37900
379000
MIN
or
MAX
mm
max
mm
max
mm
max
1978
COSTS
$
130
230
600
1,000
2,800
4,300
,000
,000
,000
,000
,000
,000
INFLATION
COSTS
19
34
132
220
616
946
$
,500
,500
,000
,000
,000
,000
DESIGN
COSTS
$
47
84
234
390
1,093
1,679
,800
,700
,000
,000
,000
,000
1981
COSTS
$
197
349
966
1,610
4,509
6,925
,300
,200
,000
,000
,000
,000
Ozonation capital costs are two to four times higher than
chlorination capital costs depending on the ozone contactor's
efficiency. The estimates in Table 34 assume the small plant will
use an air-fed ozonation system and the two larger plants will
use an oxygen-fed system. Ozone production from air reportedly
requires about fifty percent more capital than ozone production
from oxygen (Nail, 1980) per unit of ozone produced.
Amortized ozonation capital costs are shown in Table 35 for
several interest rates and amortization periods.
TABLE 35. ANNUAL OZONATION CAPITAL COSTS
INTEREST
RATE
%
0
0
3
6
9
12
15
AMORT .
PERIOD
yrs
1
20
20
20
20
20
20
MAX
or
MIN
mm
max
min
max
min
max
min
max
min
max
min
max
min
max
3790
$197,300
349,000
9,900
17,000
13,000
23,000
17,000
30,000
21,000
38,000
26,000
47,000
32,000
56,000
PLANT SIZE
m3/d
37900
$966,000
1,610,000
48,000
80,000
65,000
108,000
84,000
140,000
105,000
175,000
129,000
216,000
155,000
258,000
379000
$4,509,000
6,925,000
225,000
346,000
302,000
464,000
392,000
602,000
492,000
754,000
604,000
927,000
722,000
1,107,000
101
-------
The annual ozonation capital costs per unit of wastewater
treated are shown in Table 36. The unit costs shown are based on
an assumed interest rate of 9% and an amortization period of 20
years. The effects of underutilization on the unit costs is also
shown.
TABLE 36. OZONATION UNIT CAPITAL COSTS
PLANT ANNUAL UTILIZATION
SIZE COST
m3/d $
3790 21,000 - 38,000
37900 105,000 - 175,000
379000 492,000 - 754,000
RATE
%
100
80
60
100
80
60
100
80
60
ANNUAL UNIT
COSTS
$/l,000 m3
15.18 - 27.47
18.98 - 34.34
25.30 - 45.78
7.59 - 12.65
9.49 - 15.81
12.65 - 21.08
3.56 - 5.45
4.45 - 6.81
5.93 - 9.08
Operating and Maintenance Costs for Ozonation--
Ozone operating costs differ from those of chlorine
disinfection primarily in that the chlorine cost is replaced by
power for producing ozone. Opatken (1978) estimated power needed
for ozone production from air to be 30.8 kWh/kg, and Gumerman
(1978) estimated 16.5 kWh/kg when generating from oxygen. These
estimates are consistent with the estimates cited in the energy
discussion above. The power costs at 6 cents per kWh are shown in
Table 37 using Opatken's estimate for power utilization and
contactor efficiencies ranging from 40 to 90%. The 22 kWh/kg 03
TABLE 37. OZONATION POWER COSTS
PLANT
SIZE
m3/d
3790
37900
379000
OZONE POWER
DEMAND DEMAND
kg/d kWh/kg 03
21
209
2092
- 47
- 472
- 4722
22
3^
22
30
22
30
.8
.8
.8
10
14,
100,
141,
1,008,
1,410,
ANNUAL
COSTS
$
,100 -
160
700
000
000
000
-
-
-
- 2
- 3
22,700
31
227
318
,275
,185
,700
,500
,400
,000
,000
UNIT
COSTS
$/l,000 m3
7.32
10.24
7.32
10.19
7.32
10.19
- 16.39
- 22.92
- 16.39
- 23.02
- 16.39
- 23.02
102
-------
power demand estimate for oxygen fed ozonators is taken from the
energy discussion above because that estimate includes the power
cost of producing the oxygen.
Labor costs and supplies for ozone are somewhat less than
those for chlorine (Gumerman, 1978). The labor costs per unit of
wastewater treated are shown in Table 38. The costs shown are
based on Gumerman's hour estimates and an assumed wage rate of
$10/hour.
TABLE 38. OZONATION LABOR COSTS
PLANT
SIZE
m3/d
3790
37900
379000
ANNUAL
LABOR
hrs
550
900 - 1500
4,000 - 7,000
ANNUAL
COSTS
$
5,500
9,000 - 15,000
40,000 - 70,000
UNIT
COSTS
$/l,000
3.98
0.65 -
0.29 -
m3
1.08
0.51
Total O & M costs are dependent on the efficiency of ozone
absorption and on ozone generation; and, providing that
electrical efficiency does not deteriorate, these costs per unit
will be nearly invariant with utilization rate. Under the most
favorable circumstance, ozone 0 & M costs are 35 percent higher
than chlorine. In the worst case, ozone costs are nine times
chlorine costs. As electrical costs increase, ozone economics
will deteriorate, unless substantial improvements in ozone
generation efficiency can be made. Ozonation 0 & M costs based on
the assumptions stated above are summarized in Table 39.
TABLE 39. OZONATION O & M COSTS SUMMARY
$/l,000 m3
PLANT SIZE
m3/d
T79037900379000
Labor
Power
Supplie^
3.
10.24
1.95 -
98
- 22.92
2.82
0.65 -
7.32 -
0.96 -
1.08
16.39
1.30
0.29 -
7.32 -
0.34 -
0.51
16.39
0.59
Totals 16.17 - 29.72 8.93 - 18.77 7.95 - 17.49
103
-------
Ozonation Costs Summary—
Total costs for ozone are the sum of capital and O & M costs
and will have a wide range depending on generation efficiency,
absorption, and utilization. The ozonation costs per unit of
treated wastewater are summarized in Table 40. The values shown
are sums of the values shown in Tables 36 and 39 and are
constrained by the assumptions described in the discussion above.
TABLE 40. OZONATION COSTS SUMMARY
$71000 m3
PLANT
SIZE
m3/d
3790
37900
379000
PERCENT
60
41.47 - 75.50 35
21.58 - 39.85 18
13.88 - 26.57 12
80
.15 -
.42 -
.40 -
UTILIZATION
64
34
24
.06 31.
.58 16.
.30 11.
100
35 - 57.
52 - 31.
51 - 22.
19
42
94
Notes - absorbed ozone dosage = 5 mg/L
- minimum unit cost corresponds to 90% transfer
efficiency
- maximum unit cost corresponds to 40% transfer
efficiency
- interest rate = 9%
- amortization period = 20 years
Ozone costs for disinfection will be significantly higher
than chlorine costs. The difference will range between $8.14 and
$54.71 per 1000 cubic meters treated. Increases in energy
prices and interest rates will both increase the difference
between chlorination and ozonation costs. Some perspective can be
gained by comparing the difference between ozone and chlorine
disinfection with the total cost of wastewater treatment which is
assumed for this analysis to be about $430 per 1000 m3 in a new
system. Hence, the choice of ozonation over chlorination will
increase total wastewater treatment costs between 2 and 13%.
104
-------
SECTION 9
RISK MODEL
One objective of this risk analysis project is to develop a
risk model from the data base gathered on the risks and benefits
associated with the various wastewater disinfectants. The outputs
of the risk model serve as the inputs to the decision making
process and should be easily interpreted by the decision maker.
The model should address, in as quantitative manner as possible,
the relative risks and benefits for the disinfectants chlorine,
chlorine-dechlorination, ozone, and ultraviolet radiation as well
as the no disinfection alternative.
The risk model for this project is based solely on the
historical probability of certain events occurring. The health
and environment risks associated with the transportation and use
of the various disinfectants have been discussed and quantified
in the previous sections of the report. This section will present
two examples of how this quantitative information can be used by
the decision maker in comparing disinfection alternatives.
RISK MODEL DATA BASE
The total reliance on historical data for this risk
assessment represents a limitation, and in the case of some
disinfectants a severe limitation, to the development of
reliable, quantitative information for the decision maker. The
risk assessment output can be no better than the quality of the
published data and any assumptions required to quantitate a risk.
Nevertheless, historical data is often the best source of
reliable data, and its sole use does not preclude the development
of reliable, quantitative information.
A subjective estimate of the quality of the historical risk
assessment data base for the various wastewater disinfectants is
presented in Table 41. The ratings are boiled on both the quantity
of published data and their quality. For- those data sections
classified as insufficient (3) or not applicable (4) a
quantitative comparison of risks is not possible.
105
-------
TABLE 41. ESTIMATE OF HISTORICAL DATA QUALITY
No Chlorine Ultra-
Disinf. Chlorine Dechlor. Ozone violet
Risks
On-site Use 4
Transportation 4
Toxicological effects
of residual 4
Formation of Hazardous
Reaction Products 4
Human Health Risks 3
Evironmental Risks 3
Economic Analysis
Cost Factors 1
Energy Use 4
Ratings :
1 sufficient data
2 sufficient data
2
1
1
1
1
1
1
1
only
3
1
3
3
3
3
1
1
from parallel
3
4
3
3
1
3
1
1
industry
3
4
3
3
3
3
3
1
3 insufficient data
4 not applicable
The estimate of historical data quality in Table 41 clearly shows
the predominance of data on only one disinfection alternative,
ch1orination. This imbalance in the data base causes chlorination
to receive excessive attention in risk assessments drawn from the
data base which- makes objective assessments difficult to achieve.
RISK MODEL EXAMPLES
Two examples will be presented which illustrate the
quantitative aspects of the historical data presented in previous
sections. The two examples were chosen to represent maximum and
minimum treatment plant capacities with Example A being the small
plant. The alternative disinfection processes are limited to
chlorination and ozonation in these examples; however, the other
alternatives can be incorporated in similar risk analyses in the
same manner as demonstrated herein.
The uncertainties associated with the calculations are
difficult to quantify but are highly dependent on the assumptions
used for such variable factors as applied dosages, size of total
workforce, the local ecological community and on the quality of
the historical data as discussed above.
106
-------
Example A
A wastewater treatment plant with a capacity of 3,790 cubic
meters per day discharges to a flowing freshwater stream. The
total workforce for the treatment facility is 6 full-time
employees. The assumed chlorine dosage is 6 mg/L, and the
alternative ozone dosage is 5 mg/L with an 85% mass transfer
efficiency using air as the feed gas.
Transportation Risks--
Chlorine is to be obtained from a manufacturer located 644
kilometers from the treatment plant. The yearly quantity of
chlorine required can be calculated from the applied dosage rate.
Annual chlorine use = 6 mg/L x 3,790,000 L/d x 365 days/yr
x 10~6 kg/mg
= 8300 kg/yr
Two examples for transportation are presented to estimate
the sensitivity of the transportation mode.
Case' 1. Chlorine is transported by truck using 114 kg cylinders
the entire 644 kilometers producing an annual haulage of
5,345 metric ton-km. Multiplying this annual haulage by
the accident factors contained in Table 9 produces the
risk estimates shown below.
deaths - 0
injuries - 0.014/yr
property damage - $1.92/yr
releases - 0.64 kg/yr
Case 2. The chlorine is transported 604 km by railroad then 40
km by truck using 0.91 metric ton containers. This
corresponds to 5013 metric ton-km by railroad plus 332
metric ton-km by truck. The risk estimates for this mode
of transportation are calculated as above and are shown
below.
deaths - 0
injuries - 0.000075/yr
property damage - $0.31/yr
releases - 0.17 kg/yr
The above examples indicate, as does Table 9, that truck
transportation of chlorine in 114 kg cylinders has a
significantly higher accident rate than other modes of
transportation. However, relatively small shipments of chlorine
using the small cylinders, as in the above example, do not pose a
significant risk for human health and property damage. There is a
low sensitivity for transportation mode when dealing with small
quantities of chlorine.
107
-------
On-Site Accidents (Chlorine)—
The on-site accidents information is contained in Tables 2
and 4, and the chlorination risk analysis procedure is described
in Section 4. The severity rate for all accidents is calculated
as:
Lost work time = 6 employees x 1985 man-hr/employee/yr x
210.9 x 10~6 lost work days/man-hr
= 2.5 lost work days/yr.
The 2.5 lost work day severity rate represents all accident
types. To obtain the lost work days from exposure to chlorine the
total figure is multiplied by the assumed rate of 4% giving 0.1
lost work days per year from chlorine exposure. This last
calculation is based on the assumption that lost work time is
about the same for each type of accident.
Energy Use—
Chlorine.— The total kilowatt-hours required per year is
calculated using an efficiency factor of 3.7 kWh per kg Cl2 from
Section 8.
Chlorine energy use = 8300 kg/yr x 3.7 kWh/kg
= 31,000 kWh/yr.
Ozone.—The energy use data for ozone are contained in
Section 8 and Table 19. The total ozone required per day is:
Ozone use = 5mg transf erred/L x 1 mg appl ied/0_.85 mg
transferred x 3,790,000 L/day x 10~6 kg/mg
.= 22 kg applied/day-
The total kilowatt-hours required per year is calculated using an
efficiency factor of 30 kWh/kg ozone. This factor represents an
average value since Section 8 data do not clearly indicate any
economy of scale.
Ozone Energy Use = 22 kg/day x 365 days/yr x 30 kWh/kg
= 240,000 kWh/yr.
Cost —
The cost data in Section 8 include a variety of assumptions
on utilization rates, inflation, interest rates, etc. and the
user is directed to the section for details. The calculated costs
represent the sum of capital plus operation and maintenance.
108
-------
Chlorine.—The cost data for chlorine disinfection are
summarized in Table 32. With a dosage of 6 mg/L and a utilization
of 80%, the total chlorine disinfection cost is calculated as:
Chlorine Cost = $15.23/1000 m3 x 3.790 (1000 m3)/day x
365 days/yr
= $21,000 /yr.
Ozone.—The cost data for ozone are summarized in Tables 36,
37, 39, and 40. The unit capital costs shown in Table 36 are for
40 and 90% mass transfer efficiency. The unit capital cost for an
85% efficient plant is calculated by interpolation as:
Ozone Unit Capital Cost = $18.98 + ((34.34 - 18.98) x
( (90-85)/(90-40)))
= $20.52/ 1000 m3
The power cost is calculated from the energy calculation above.
Ozone Power Cost = 240,000 kWh/yr x $0.06 /kWh
= $14,400 /yr.
Using unit costs for labor and supplies as shown in Table 39, the
ozone costs are calculated as:
Ozone Costs = ($20.52 + 3.98 + 1.95)/1000 m3
x 3.79 (1000 m3)/d x 365 d/yr + $14,400
= $51,000 /yr.
Ecological Effects—
The ecological effects of the facility in Example A can be
estimated if the following assumptions are accepted.
1. Aquatic organisms do not avoid the effluent, i.e., fish
do not swim downstream to lower concentrations.
2. Synergistic effects with other wastewater constituents
are not present.
3. The biological community is represented adequately by
Bluegill, Channel catfish, Rainbow Trout, Brown Trout,
Daphnia magna, and an oligochaete worm.
4. Water quality parameters and experimental protocols in
the literature are representative of a field situation.
5. Non-lethal concentrations are determined to be one-half
the LC-50 concentrations unless species specific data
are available (Seegert and Bogardus, 1980).
109
-------
6. TRC removal mechanisms and dilution in a flowing stream
can be estimated by C = Co exp(-kt) where C is stream
concentration, Co is initial (diluted) stream
concentration at the outfall, t is residence time in the
stream, and k is a rate constant.
7. Stream velocity is 3 km/hr, and essentially all TRC is
removed or diluted within 8 km from the outfall.
8. Outfall TRC concentration is 0.5 mg/L, which is diluted
by the flowing stream at the outfall in a ratio of 1:10.
The summary of literature (Appendix II) on TRC toxicology
for the representative aquatic community assumed above is shown
in Figure 2, Section 4. The data are very scattered, and as such
do not provide a basis for a dose-response function. In
particular, the magnitude of a threshold TRC concentration is not
consistent thoughout the literature. Sub-lethal effects data are
not shown due to their inconsistencies relative to apparently
lethal concentrations. Segregation of the data by exposure time
or temperature does not significantly decrease variability.
However, the assumption of a threshold at 0.5 of LC-50
concentrations provides an approximate method to determine
relative population stress on a species basis.
Population stress was determined in a semi-quantitative
manner by measuring the graphical distance between the local
stream concentration and the bulk of mortality data points in
Figure 2. Large graphical distances above local stream
concentrations indicated low stress. Small graphical distances
above local stream concentrations indicate moderate to high
stress. Large distances below local stream concentrations
indicate extremely high stress. Equating relative mortality to
population stress avoids the misconception of massive aquatic
kills. In a chronic situation, populations in high TRC areas will
be permanently depressed, and will be comprised of individuals
with high resistance to TRC.
Areas of moderate TRC concentrations will have aquatic
populations under more moderate stress, including sub-lethal
effects such as decreased activity, decreased reproduction, and
irritability. However, even in areas of moderate TRC
concentrations, weakened individuals will show increased
mortality -
Table 42 ta!u3ates values of population stress at various
distances downstream from the Example A facility outfall. These
values show thac trout populations near the outfall will be
significantly depressed. Channel catfish and D. Magna populations
will have mostly sub-lethal effects near the outfall. Bluegill
and Oligochaete worm populations will be unaffected.
110
-------
Human risk from disinfection of wastewater almost wholly
results from potential ingestion. Although it is unreasonable to
project direct consumption of water from the outfall of a
disinfection facility, it is possible to examine human risk from
the proportion of disinfected wastewater which is ingested from
public drinking water. Adequate data exist for estimation of
human risk from certain carcinogenic products in chlorinated
wastewater given that certain bounding assumptions are made.
TABLE 42. EFFECTS OF EXAMPLE A FACILITY ON AQUATIC ORGANISMS
Dist.
Down-
stream
km
0
0.5
1
1.5
2
3
4
6
Cone.
mg/1
0.050
0.034
0.023
0.016
0.010
0.005
0.002
0.0005
Blue-
gill
0
0
0
0
0
0
0
0
Ratings :
Species
Channel Rainbow
Catfish Trout
1
0
0
0
0
0
0
0
0 =
1 =
2 =
3 =
4 =
3
2
2
1
1
1
1
0
no stress
light stress
Brook
Trout
3
2
2
1
0 v
0
0
0
D. Magna
1
1
1
1
1
1
1
0
Oligo-
chate
0
0
0
0
0
0
0
0
moderate stress
high stress
extreme stress
These assumptions are as follows:
1. the risk from consumption of chlorinated wastewater is
represented by the risk of ingestion of chloroform and
trichloroethylene (other compounds do not have an
adequate data base from which to estimate risk),
2. the proportion of chloroform and trichloroethylene to
chlorine dose is 0.01 (extrapolated form Naek and Doerr,
1978; Hoehn, et al., 1978),
3. the proportion of chlorinated wastewater entering
drinking water systems will not exceed 20%,
111
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4. the human health risk from ingestion of chloroform
ranges from 0.37 cancer cases/million people/ug/L (Neal,
1980) to 1.7 cancer cases/million people/ug/L (NAS,
1977) ,
5. the human health risk from ingestion of
trichloroethylene is 0.11 cancer cases/million
people/ug/L (NAS, 1977) to 0.13 cancer cases/million
people/ug/L (Neal, 1980).
If these assumptions are valid, human risk can be bounded between
zero (no wastewater in drinking water) and some value which is an
upper bound estimate of human risk.
For the Example A, chlorine dosage of 6 mg/L, chloroform
and trichloroethylene concentrations are both 0.06 mg/L (60 ug/L)
at the outfall and are diluted to 6 ug/L by the flowing stream.
Using the upper bound cancer risk of 1.7
cancer cases/million people/ug/L and an upper bound percent
wastewater in drinking water of 20%, the lifetime risk from
chloroform originating •pi"om H-e facility in Example A is:
R (Chloroform) = 1.7 cancer cases/million people/ug/L x 6 ug/L x
0.2 L wastewater/L water consumed
= 2 cancer cases/million people.
Lifetime risk from trichloroethylene originating from the
facility in Example A is
R (Trichloroethylene) = 0.1? cancer cases/million people/ug/L x
6 ug/L x 0.2 L wastewater/L water consumed
R =0.16 cancer cases/million people.
or
R = 0.00000016 cancer cases/1 ifetime
Thus, total human health risk from wastewater qhlorination in the
facility of Example A ranges from 0 to 2 x 10". It is rather
unlikely that 20% of a given drinking water supply will be
chlorinated wastewater, since Swayne, et al. (1979) showed that
almost half of the population sampled (80 million) ingests
drinking water with no chlorinated wastewater, and only 0.7
percent have wastewater concentrations above 5 percent during
average flow conditions. However, 2 x 10~6 represents an upper
bound of risk.
Comparative risk from ozonolysis by-products cannot be
estimated due to a lack of data.
Human health risk resulting from recreational use of the
stream in Example A can be calculated using the Cabelli model
presented in Section 7. Assume k is 0.15 per day and, from the
information given above, Qe/(Qe + Qs) is 0.1. Also assume the
112
-------
enterococci concentration of the undisinfected wastewater
discharge is 10,000/100 ml. The human health risk from swimming
in the receiving stream at a point two days travel distance
downstream would then be:
y = 11.85 log(10,000 x 0.1 x exp(-0.15 x 2)) - 0.58
= 33 cases of gastrointestinal distress/1000 swimmers.,
Summary—
Table 43 summarizes the risks associated with chlorination
and ozonolysis for Example A. Although total disinfection costs
and energy use are higher for ozone, human health and
environmental risk for ozonolysis are much lower than for
chlorination.
TABLE 43. RISKS SUMMARY - EXAMPLE A
Description
Chlorination
Ozonation
Transportation
Case 1 - truck only
114 kg cylinders
Deaths/yr
Injuries/yr
Property damage - $/yr
Releases - kg/yr
Case 2 - rail + truck
0.91 metric ton cylinders
Deaths/yr
Injuries/yr
Property damage - $/yr
Releases - kg/yr
On-site Accidents - lost work days/yr
Energy Use - kWh/yr
Cost - $/yr
Human Health Risk -
Cancer cases/lifetime
Ecosystem Effects
0
0.014
$1.92
0.64
0
0.000075
$0.31
0.17
0.1
31,000
21,000
0 - 2x10
-6
trout population
stress near the
outfall
not
applic-
able
Insuff.
Data
240,000
51,000
Insuff.
Data
None
Example B
This wastewater treatment plant with a capacity of 379,000
cubic meters per day is also located on a flowing freshwater
stream. The total workforce is 150 full-time employees. The
assumed chlorine dosage is 10 mg/L and the absorbed ozone dosage
113
-------
is 5 mg/L with a 90% mass transfer efficiency using oxygen as the
feed gas.
Transportation—
Chlorine is to be obtained from a manufacturer located 965
kilometers from the treatment plant. The yearly quantity of chlorine
is calculated as follows.
Chlorine Use/yr = 10 mg/L x 379 x 106 L/day x 365 days/yr
x 10~9 metric tons/mg
= 1,380 metric tons/yr
Annual Haulage = 1,380 metric tons/yr x 965 km
= 1,332,000 metric ton-km/yr
Chlorine for large plants is typically transported by
railroad tank cars. The probable risks and damages arising from
this chlorine usage are calculated using the factors obtained
from Table 9. The resulting estimates are shown below.
Deaths - 0.0004/yr
Injuries - 0.02/yr
Property damage - $79.92/yr
Releases - 44 kg/yr
On-Site Accidents--
Chlorine.--The on-site accidents data are contained in
Tables 2 and 4. The lost work day severity rate for 150 employees
is calculated as follows.
lost workdays/yr = 150 employees x 1958 man-hr/employee
x 749.3 x 10~6 lost work days/man-hr
x .04 chlorine lost days/lostwork days
= 8.8 lost work days/yr
Energy Use--
Chlorine.—The total kilowatt-hours required per year
are calculated using a value of 1.41 kWh/kg C12 as discussed in
Section 8.
Chlorine energy use = 1,380,000 kg/yr x 1.41 kWh/kg
= 1,950,000 kWh/yr
Ozone.—The ozone energy utilization rate is assumed to be
22 kWh/kg based on the discussion of oxygen fed ozonators in
Section 8.
Ozone production = 5 mg absorbed/L x 1 mg applied/0.9 mg
absorbed x 379 x 106 L/day
x 365 days/yr x 10"^ metric tons/mg
= 768 metric tons/yr
114
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Ozone energy use = 768 metric tons/yr x 22 kWh/kg
x 1000 kg/metric ton
= 16,900,000 kWh/yr
Cost—
Chlorine.—A unit chlorination cost of $4.98/1000 m3
corresponding to a dosage of 10 mg/L and a utilization rate of
80% was chosed from Table 32 for this example.
Chlorination cost = $4.98/1000 m3 x 379 (1000 m3)/day
x 365 days/yr
= $690,000/yr.
Ozone—A unit ozonation cost of $12.40/1000 m3 corresponding
to an absorbed dose of 5 mg/1, a transfer efficiency of 90%, and
a utilization rate of 80% was chosen from Table 40 for this
example.
Ozonation cost = $12.40/1000 m3 x 379 (1000 m3)/day
x 365 days/yr
= $1,720,000/yr
Ecological Effects--
The ecological effects of the facility in Example B can be
estimated using the same assumptions discussed for Example A.
Example B, however, is assumed to be diluted to a ratio of 1:3
with an undiluted TRC of 1 mg/L at the outfall. These TRC levels
account for the much larger outfall volume and chlorination which
occurs in a large wastewater facility.
Table 44 shows population stress values for various
distances downstream of the Example B facility. All aquatic
organisms are stressed near the outfall. Daphnia magna populations
may be non-existent near the outfall, which may cause additional
stress to predator populations due to lack of food organisms.
Channel catfish and Oligochaete worm populations will be highly
stressed, and Bluegill populations will be moderately stressed
near the outfall.
Human Risk--
The human risk from a facility described as Example B is
calculated as in Example A. For Example B with chlorine dosage
equal 10 mg/L the chloroform and trichloroethylene concentrations
are 100 ug/L. This is diluted to 33 ug/L by the flowing stream at
the outfall. Lifetime cancer risks are shown be1ow.
R(Chloroform) = .00000017 cancer cases/ug/L/lifetime
x 33 ug/L x .20 L wastewater/L water
R = 0.000011 cancer cases/lifetime
115
-------
TABLE 44. EFFECTS OF EXAMPLE B FACILITY ON AQUATIC ORGANISMS
Species
Dist.
Down-
stream
km
0
0.5
I
1.5
2
3
4
6
8
Cone.
Blue-
gill
Channel Rainbow Brook
Catfish Trout Trout
D. Magna
Oligo-
chate
mg/ 1
0.
0.
0.
0.
0.
0.
0.
0.
ni
33
23
15
10
07
03
01
003
1
2
2
1
1
0
0
0
0
0
Ratings :
3
3
3
2
1
0
0
0
0
0 =
1 =
2 =
3 =
4 =
4
4
3
3
3
2
1
0
0
no stress
light stress
moderate stress
high stress
extreme stress
4
4
3
3
3
2
1
0
0
4
4
3
2
1
1
1
1
0
3
2
1
1
0
0
0
0
0
R(Trichloroethylene) = 0.00000013 cancer cases/ug/L/lifetime
x 33 ug/L x .20 L wastewater/L water
R = 0.00000086 cancer cases/lifetime
Thus, total human health risk from wastewater chlorination in the
facility of Example B ranges from 0 to 0.000011. As in Example A,
it is very unlikely that 20% of a given drinking water supply
will be chlorinated wastewater (Swayne, et al., 1979), since only
0.7% of the population is exposed to drinking water containing
over 5 percent chlorinated wastewater during average flow
conditions. Lifetime cancer risk of 0.000011 is an upperbound
estimate.
As in Example A, comparative risk for ozonation by-products
cannot be estimated due to lack of data.
Human health risk from swimming in hr receiving stream for
Example B can be calculated using the s^.me assumed values as in
Example A except that Qe/(Qe + Qs) is now 0.33. Therefore,
y = 11.85 log(10,000 x 0.33 x exp(-0.15 x 2)) - 0.58
= 40 cases of gastrointestinal illness/1000 swimmers.
116
-------
Summary—
Table 45 summarizes the risks associated with chlorination
and ozonalysis for Example B. As was found for Example A, human
and environmental risks of ozonalysis is much lower than for
chlorination. Energy use and total disinfection costs are still
higher for ozonation than for chlorination even though the
assumptions selected for Example B extremely favorable to
ozonation and somewhat unfavorable to chlorination.
TABLE 45. RISKS SUMMARY - EXAMPLE B
Description Chlorination Ozonation
Transportation
rail tank cars
Deaths/yr 0.0004 not
Injuries/yr 0.02 applic-
Property damage - $/yr $79.92 able
Releases - kg/yr 44
On-site Accidents -
lost work days/yr 8.8 Insuff.
Data
Energy Use - kWh/yr 1,950,000 16,900,000
Cost - $/yr 690,000 1,720,000
Human Health Risk -
cancer cases/lifetime 0 - 12xlO~° Insuff.
Data
Ecosystem Effects all organisms None
highly stressed
near outfall
117
-------
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21, 1980.
Waugh, G.D., "Observations on the Effects of Chlorine on the
Larvae of Oysters (Ostrea edulis L.) and Barnacles (Elminius
modestus (Darwin)," Ann. Appl. Biol., 54, 423, 1964.
Wedemeyer, G.A., Nelson,Nancy C., and Yasutake, William T.,
"Physiological and Biochemical Aspects of Ozone Toxicity to
Rainbow Trout (Salmo gairdneri)," Jour. Fish Res. Bd. Can.,
36, 605, 1979.
148
-------
Weinbach, E.G. and J. Garbus, "The Interaction of Uncoupling
Phenols with Mitochondria and with Mitochondrial Protein,"
Jour. Biol. Chem. 210:1811, 1965.
Weller, R.W. and Crellin, A.J., "Pulmonary Granulomatosis
Following Extensive Use of Paradichlorobenzene," Arch.
Intern. Med. 91:408, 1953.
West, W.L. and Ware, S.A., "Investigation of Selected Potential
Environmental Contaminants: Halogenated Benzenes," U.S. EPA,
Washington, D.C., 1977.
White, G.C., "Handbook of Chlorination - For Potable Water,
Wastewater, Cooling Water, Industrial Processes, and
Swimming Pools," Van Nostrand Reinhold Co., New York, 1972.
White, G.C., "Current Chlorination and Dechlorination Practices
in the Treatment of Potable Water, Wastewater and Cooling
Water," in Water Chlorination Environmental Impact and
Health Effects, R.L. Jolley, Editor, Vol. I, 1978.
White, W.R., "The Effect 'of Low-Level Chlorination on Mussels at
Poole Power Station," Central Electricity Generating Board
Rept. RD/L/N 7/66, 1966.
Wien, R., "The Toxicity of Parachlorometacresol and of
PhenyImercuric Nitrate," Quarterly Jour, and Yearbook of
Pharmacy. 12:212, 1939.
Windholz, M. (Ed.), The Merck Index, 9th ed. Merck and Co.,
Rahway, New Jersey, 1976.
Winklehaus, C., "Ozonation in a Better Perspective," JWPCF, 49,
190, 1977.
Wolf, E.G., et al., "Bioassays on the Combined Effects of
Chlorine, Heavy Metals, and Temperature on Fish and Fish
Food Organisms," Batte1le-Northwest Research Lab., draft
manuscript.
Wright, F.C., et al., "Metabolic and Residue Studies with 2-
(2,4,5-trichlorophenoxy)-ethy1 2,2-dichloropropionate , "
Jour. Agric. Food Chern. 18:845, 1970.
Zillich, J.A., "The Toxic Effects of the Grandville Wastewater
Treatment Plant Effluent to the Fathead Minnow, Pimephales
promelas, Nov. 1969.
Zimmerman, P.W., and Berg, R.O., "Effects of Chlorinated Water on
Land Plants, Aquatic Plants, and Goldfish," Contrib. Boyce
Thom£j5cm I_n£t_._, 6 , 39, 1934.
149
-------
Zondek, B. and Shapiro, B., "Fate of Halogenated Phenols in the
Organism," Biochem. Jour. 37:592, 1943.
150
-------
APPENDIX A
U.S. CHLORINE PRODUCERS AND PACKAGERS
ALPHABETICAL LISTING OF CHLORINE PRODUCERS IN THE
« Alualniai Company of AMTiaa
Paint CoMfort. T«» (lot* I)
II AMI Spaalalty Matala Corp.
(onlay. Utah
Aiarlaan Nagnaalua, Co.
3nya«r, Taiaa (Not* 8)
M BASF Uyandotta Corp.
Galaur, Louisiana
M k-Miutok Chamloal Co. Sub*.
Brunswick Pulp * Papar Co.
Bru»ulek, Gaorgla
Champion total-national Corp.
1. Canton. North Carolina
2. Houston. Taiaa
H Conant Chaailoal Company
1. Calvart City, Kanluoky
2. Con»«nt. Loulllan*
H Plaawnd Snaau-ook Corp.
1. Daar Park. Ta«aa
2. Dalawara City, Dalaiara
I. UPorta. Taiai
4. NoDlla, Alabama
5. Huael* Shoals. Alabama
H Don ChaaUoal US*
1. Fraaport, Taiaa
2. Hidland. Mlcnlgan
3. Plttsburg, California
4. Plaquamlna. Loulalana
M 1. I. «u Pont da Nawur» * Co.. In«.
1. Corpua Cbrlatl, Tuaa
2. »ia«»ra Falla. Haw tork
H Ethyl Corporation
Baton loufa. Loulalana
M FMC Corporation
South Oiarlaatoa. Haat Vlnlnla
M Foneaa Plaatloa Corp.. USA
Baton touia, LouUlana
Fort Houarit Papar Coapaay
1. Craan Bay, Maoonaln
X. Hnakotaa, OKlahou
M Ganaral Ha«trlo Co.
Nt. Varnan, Indiana
M Caorila-Paaine Corp.
1. BalllnglM, Waahlngton
2. Plaquaclna. Louisiana
M Haroulaa, Inc.
HopaMall, Vlnlnla
N Hookar Cnailoala 4 Plaatlea Corp.
1. nontaiua, HloHlian
2. Niagara Falla. Na« Tork
3. Tacou, Waahlngton
4. Taft, Loulilana
Hookar - IHC Joint Vantura
Falla, Hav lork
H Int'l Mlnarals 4 Chamlaal Corp.
1. Aantabula, Ohio
2. Orrington. Main*
H Kalaar Alualnui 4 Chamiaal Corp.
Cramarcy, Louisiana
H Undan Chamlcala A Plaatloa, Inc.
1. Aoma. North Carolina
2. Brunawick, Caorgla
3. Undan, Naw Jaraay
4. taundsvilla. Wast Virginia
5. Syraeuaa, Naw Tork
H hobay Chamiaal Corp.
Baytown. Taias
II Monsanto Company
Saugat, Illinois
N 01in Corporation
1. Auguata, Caorgla
2. Charleston, Tannaaaaa
3. Holntoah. Alaoama
4. Niagara Falla. Naw Tork
dragon Hatallurgloal Corp.
Albany, Dragon
M Pannualt Corporation
1. Calv.rt City. Kantuoky
2. Portland. Dragon
3. Taooma, Washington
4. Wyandotta, Miohltan
M PPG IMuatrlaa, Ine.
1. BarMrton, Onlo
2. Laka Charlai, Louisiana
3. Nau Kartlnavllla. U. Virginia
M IMI Coaipany*
Aahtabula, Ohio
N Shall QiMloal Coapany
Oaar Park, Taiaa
M Stauffar Chaalcal Coapany
1. Handaraon, Navadi
2. Laltoyna. Alabaaia
3. St. Gaorlal. Louisiana
Tltanlua Matala Corp. of Aaarlca"
T1MET Dlvn.
Handaraon. Navada
Vartao Chaailoal Coapany
Vickaourg. Hlaalaalppl
M Vulcan Hatarlals Company.
O>mleals Division
I. D«nv»' City. T«««s
2. GalaBar, Louisiana
3. Port Edwards, Wisconsin
H. Uienita, Kanaas
H Wayarbaausar Coapany
Longvlaw, Washington
• Joint aub»idl.r» of National Dlatlllara 4 Chaailoal Corp. and U.S. ^••^
aodlu. 4 ohlorlha produot. »ld By U.S. loduatrlal Ch-loal. Co.. Division
of National Dlatlllara.
aa Tltanlua Matala Corp. of Aa>arUa la a Joint aubaldlary of NL Induatriaa,
. a*d Allaghaoy Ludliai Corp.
Reproduced from
best available copy.
151
-------
U. S. CHLORINE PRODUCERS BY STATES
Mle
e.1.
. Ice.
art
C.
•ta la »*rla(*
»>ae>ee
Mr
(•IM Oil
e T
C T
tea Hater* t Mfere ttn.
Km OMieele. toe.
Stem Caemlal to.
••ami*Ul>i XeMe,
Ulteil WltereeJ tor*.
t T
C T
et
C T
C T
C T
C T
« t
tlleiit
7
MlaMre .
•Heave Cltr
fterUe
n.
Uere
at. Petereevri;
It. aetereMri
TMM
1*
SWlMe
UP*ru
terra Mute
•urllMtM
KM*a*
Klemu
leetuelfT
Leulaien*
C*lM*r
•tear**
IX. Ce»rlel
•arrUne
•alt lea**
eneeigen
irraneett.*
tlllM
allle' —l.*r»*l C ,.
J*)M* OMMleele. toe.
• * a 3 OMleal to., tec.
TManii a«!.•!•* raaaml to.
. to.
i OtMleal to.
LlaM CMeueel to.
Cere.
I.I. 3»*el Ouaieeli IM.
Jea** Oiaeleela. toe.
Ulrlea OtMleel toe.
lewtll* Meter Lea*., toe.
aaa Mtariel* to. (Mta* F.2)
eilcel*. toe.
•ft. to.
*. toe.
e t
e T
C T
t t
c t
C T
C T
C T
C T
C T
C -
C t
- T
C T
C t
C T
C -
- I
C I
C T
C T
C T
C T
Mileel*. toe.
MkleM CMeleal to.
tauaaa 6I«*
•reler. toe.
J*M* tl 1-iU. toe.
rejr*laae mi*eer CHaailael to.
Tueaiini miinr« CkMleel to.
ftvvtb tov^lLft*
ftaaml to., be.
ill torperrtlM (Mta •>
mltM CbMleel toe.
IJkltae ChMlcel toe.
IklLae CaaalMj toe.
Hrv OhMteel to.
topu. CkriMI
Mile*
C aaea
Fart North
eauaton
UPort*
Mlt Leea Cltr
lelt Uat Cltr
• It! Oaaeleel to.
Carrell ChM. t Crreteiue* toe.
MarlCa*. toe.
IhMeeea-Mrwe' CMaueal to.
aMrlBe*. toe.
MerlC**, toe.
aeuu Teie* CUeriM. toe.
J.It. Jane* CbMlctl to.
Dili* ChMle*! to.
MerlCeM. toe.
Tut«Mr ChaMleel to.
aaaatcn fhaaieal to.
••* Dkaalael*. toe.
•lniar> CMalael to.
Itatf Caaejleel toe.
C T
C T
C T
C T
e T
e T
c T
C T
C T
c t
C T
11
t T
C T
C T
C T
C T
e T
C T
e T
e T
C T
C T
C T
C T
C T
C T
C T
C T
C T
C -
C T
HeMfllt* Cheelc*: to. V*;-T*I Dlv.
Van Hater* i ae«*r* Olvft.
C T
C T
Jen** Cheelcal*. toe.
Paanitelt toraeratlen (MM P)
C T
t '
tliaaikiippl
«lcea»i*r§
*. lac.
ol to.
C T
C T
a.. «.-iir,«»lU« Pit] ChMleel to., toe.
It. tlkar.* P I t 1 ChMleel to., toe.
Jeaa* Chaalcel*. toe.
•raMta CMalc*! to.
(t * tor. contalnrrr; c « cyllno«.iji
c T
c T
c i
C T
. Reproduced from
[beit available copy
152
-------
APPENDIX B
SUMMARY OF REPORTED CHLORINE EFFECTS ON FRESHWATER ORGANISMS
Species
(Plants)
Chlorophyta
Algae
Pyrenoidosa
Chlorella var.
Ob.
Scenedesmus
Scenedesmus Sp .
Chrysophyta
_ , Parv.
uompnoneioa
Palea
Nitzschia
Cyanophyta
Cylindrospenc:
Aeru.
Microcystis
Miscell-Phytoplants
(Invertebrates)
Water louse
Protozoa
Tvclops. Sp.
End point
DG-50
DG-50
DG
DG
MI
DC
DG
DG
DG
Stops
Growth
No Repro
Some
Mortality
Sotnt
Mortality
Exposure Temp .
lice of Test
Kin. Solution
1,440 —
300 —
4,3207
4,3207
5,760
-.2207
4,3207 —
4,3207 —
-.3207
KG
60
L7r.ir.
30
Cone. Reference
(FRC.CRC) Kott, et •!., 1966
0.18
(FRC.CRCO ° Sf
0.4
(Ca hypo.) Palner and Maloney,
2.0 1955
(Ca hypo.) Palmer ar.d Kaloney.
2.0 1955
(CRC) Bringhar and Kuhn,
10.0 1959
Palner and Maloney,
2 1955
Palirer and Kaloney,
2 1955
Palicer and Maloney,
* 1955
., Palmer and Maloney,
1955
Brook and Baker, (1972)
0.4
Holland. C.J.. 1956
(TRO)
0.5
Ranpanathan and Small,
^^^ 1 Q 7 0
2-6
(TRC) Adars, B.A., (1927)
1.0
Reproduced /rom gtg
best available
153
-------
""*., .icies
Water Flea
Daphnla Bagna
Hater Flea
Daphnla Bagna
Hater Flea
Daphnla Bagna
Hater Flea
Daphnla magna
Hater Flea
Daphnla Bagna
KT •
DC "
HE •
D\ -
Species
Invertebrates
Scud
Scud
Scud
Scud
Scud
Scud
Scud
S<
ladpolnt Exposure
TiB*
Kin.
MT 2.880
LC-1— 240
D Repro. 20,160
LC-100 4,320
Some 60
Mortality
Mortality Threshold •—
Decreased Crovth
No Effect
(Decreased Photosynthesis)
Endpt. (Kin.) Tcap
Exposure °C
Time
LC-50 2,880 15°
LC-80 151,200 —
' LC-50 5,760 —
02 Repo 151,200 —
D Survival 161,280 —
D Repo 201.600 —
HE 43,200 —
LC-50 1,440 —
Orlonectes Virilis
Crayfish LC-50 10,080 —
T««p.
of Test
Solution
—
—
—
—
—
HC -
TRC-
FRC"
MCA
CRC
(Type)
Cone.
Bg/L
(TRC)/
0.023
(CRC)
0.035
(CRC)
0.22
(CRC)
.0034
(TRC)
.054
(TRC)
0.019
(TRC)
0.135
(TRC)
0.900
(TRC)
0.780
Cone. Reference
*
(CRC) Brlngman, C., and
4.0 Kuhn. 1959
0.125 Brungs. V.A.. 1973
Arthur, J.V., et si
, . Ellis, K.v., (1937)
(TRC)
0.5
Adams, B.A.. (1927)
(TRC)
0.5
Mot Given
Total Residual Chlorine
Free Residual Chlorine
• MonochlOTnine
- MCA + di end tri
Teat Reference " .,
Type -v-
FT Cregg. B.C.. 1974
•nd
IT Arthur, Eaton, 1971
•nd
PT Arthur, Eaton, 1971
•nd
Arthur, Eaton, 1975
FT
?
Arthur, J.V., et al.
n 1974
Arthur, J.V.. et al.
17 1974
Arthur, J.V., «t •!.
17 1974
Arthur, J.U., et al.
n 1974
Arthur, J.V., et al.
17 1974'
Arthropods -Insects
Mayfly
Centroptium, Sp.
Mayfly
ChironoBus Sp.
Hldge larvae
LC-50 480 25°
LC-50 1,440 6°
LC-80 1.440 —
0.502
(TRC)
0.071
(TRC)
7.0
n Cregg. B.C.. 1974
n Cregg. B.C.. 1974
Buctmtan. V.. 1933
154
-------
Kp«cicB Xadpt.
480 25°
.080 —
,880 15°
35 —
34 —
30 —
30 —
120 —
150 —
90 —
(Typ«>
Cone.
•g/L
(TRC)
0.027
(TRC)
0.046
0.0093
0.396
0.55
0.020
(FRC)
1.0
(TRL)
1.0
(FRC)
0.5
(FRC)
91.0
13.0
(TRC)
20.0
3.0
TMC -
Type
FT
FT
FT
FT
FT
FT
FT & S
S
S
S
S
S
S
Cregg, B.C.
Cregg, B.C.
Cregg, B.C.
Gregg, B.C.
Gregg, B.C.,
Cregg, B.C..
, 1974
i
i
, 1974
. 1974
, 1974
(1974)
1974
Learner and Edwarda
1963
Collins, J.S
Hart, K.M.,
Chang et al.
Chang et al.
Collins. J.S
Collins, J.S
.. 1958
1957
1960
1960
., 1958
.. 1958
155
-------
Species Endpolnt
v__
Molluscs
Operculate snail LC-50
Operculate snail LC-50
(Ancalosa) LC-50
«
fulaonate snail LC-50
P. heterostropha
LC-50
/
Vertebrate Anlcalc
Tadpole LC-100
Fish
Gizzard Shad MT
Coho salmon MT
Coho salnon MT
" ^o salmon —
Coho aalaon "
Coho salmon "
Coho salmon "
Coho salmon "
Cobo salmon "
Coho salmon **
Coho salmor "
Coho salnon "
Coho salvor. "
Coho salmon "
(Alevin)
Coho salmon "
Coho salmon (Fry) "
Coho salmon "
(Juvenile)
\
Uji salmon "
(Mln.)
Exposure
TIM
20,160
5,760
5,760
20,160
5,760
510
10
1,440
5.760
—
H.
n
8.640
M
5.760
n
4,320
5,760
n
n
30
30
30
30
Tcflp*
°c
_
25°
25»
j,
25°
—
--
—
__
6°
12°
12°
6°
_
6°
. 6°
10-15°
12°
9.9-10.2°
10.3°
10.3°
' 10.5-
15.0°
20.1°
(Typ«)
.Cone.
•g/L
(TRC)
> 0.810
(TRC)
0.044
(TRC)
0.086
(TRC)
> 0.810
(TRC)
0.258
2.4
0.62
0.016
0.004
_
0.68
0.64
0.55
0.54
0.89
0.677
0.181
(CRC)
0.060
0.289
(TRC)
.080-0
0.56
(TRC) ,
0.079
(TRC)
0.053-
0.082
0.29
Test
Typ«
FT
FT
FT
FT
FT
S
s
—
—
__
_
_
_
_
FT
FT
FT
FT
S
FT
.083
S
S
S
s
Reference
Arthur. J.V., et «1.
1974
Gregg. B.C.. 1974
Cregg. B.C., 1974
Arthur, J.W., et al.
i m j.
19/4
Gregg, B.C., 1974
Falnkkar, B.H.. 1960
Truchan. J.C., and
Baach. R.E., 1971
Rosenberger. D.R.,
1972
n •* .«•
_,__
Heath. A.C., 1978
Heath. A.G., 1978
Heath. A.G.. 1978
Heath, A.C., 1978
Ward. R.W.. and
DeGrave. G.M., (1978)
Heath, A.C., (1977)
Heath, A.C.. 1977
Larson, C.L., et al.
1977
Marking, L.L., and
Bills, T.D., 1977
Larson. C.L., et al.
1978 i
Seegert. C.L., and
Brooke, A.S.. 1978
Larson. C.L., et al.
1978
Larson. G.L., et al.
1978
Seegert, G.L., and
Brooks. A.S., 1978
1 Reproduced from
1 best available copy.
156
-------
Kp«ci*s
Brown Trout
«
n
M
«
(Fingerlings)
«
Brook Trout
Brook Trout
M If
« n
•I •!
n ii
M
•I II
"(Alevln)
"(Fry)
"(Juvenile)
Northern Pike
Crass Pickerel
White Sucker
H n
•i ii
Goldfish
fl
. Endpoint
MT
LC-50
LC-50
LC-50
LC-50
LC-50
MT
Activity
Depressed
MT
LC-100
LC-50
LC-100
LC-5Q
LC-50
LC-50
LC-50
LC-50
LC-100
LC-100
LC-100
LC-50
LC-50
LC-100
LC-100
(Min.)
Exposure
Time
1.440
120
180
360
660
90
10,080
10,080
10.080
2,880
5,760
1,440
720
5,760
5,760
5.760
5,760
1,800
60
60
720
5.7.-.0
2iO
1,4 in
Trap.
°C
__
—
—
—
—
_v
^^
—
_^
«*
—
20°
10.6-
10.8°
10.8°
11.1-
11.3°
4.5-7°
--
_ —
^_
i:°
~
__
0.05
(TRC)
0.06
(TRC)
10.0
(TRC)
0.360
(TRC)
0.012
(TRC)
0.09-.1
(TRC)
0.08/8
(TRC)
0.088-.09
(FRC)
0.70
(TRC)
1.0
(TRC)
1.0
(TSC) '
0.248
0.379
(TRC)
(FRC)
0.3
~T«st
Type
S
S/CF
S/CF
S/CF
S/CF
S
FT
FT
S
S
FT
S
FT
FT
FT
FT
FT
S/FT
S
S
FT
S
S
FT
Reference
Rush ton, W., 1921
Pike, D.J.. 1971
•i n M
ii n ti
•i ii ti
Pyle, E.A., 1960
Dandy, J.W.T., 1972
Dandy, J.V.I., 1971
Coventry, F.L., et al
1935
Coventry, F.L., et al
1935
Wolf. E.G., et al.
1O7&
A 3f • V
fielding, D.L., 1927 ,
Authur, J.V., et al.
(1974)
Thatcher, T.O., et al
1976
Larson, G.L., et al .
1978
Larson, G.L., et al.
1978
Larson, C.L., et al.
1978
Ebelinf, G., 1931
Hubbs, C.L., 1930
Fobes. R.L.. 1971
Authur, J.W., et al .
1974
Marleir.R, L.L., and
Bills. I.D., 1977
Panikkar, B.V., 19*0
Me Caul ey, R.V.. anH
Srntr n r i ntn
157
-------
ts«cl«s
Goldfish
Goldfish
Goldfish
Goldfish
Goldfish
Goldfish
Goldfish
Carp
n
M
••
ti
"
««
it
Golden Shiner
Golden Shiner
Golden Shiner
Golden Shiner
•Voider. Shiner
Endpolnt
MT
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-80
HT
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
Death
LC-100
LC-50
LC-sn
LC-50
Odn.)
Exposure
TIM
480
1,440
2,880
4.320
5,760
5,760
5,760
6.000
65
5,760
5,760
5,760
80
2.880
2,880
0.17
240
5,760
5,7*0
5,760
tm*>.
°C
—
20-22.5°
20-22.5°
20-22.5°
20-22.5°
12°
25°
_
—
12°
6°
6"
10°
20°
30°
—
—
6°
24°
6C
(Tyj>«)
Cone.
•8/1
(TRC)
1.0
0.49
(TRC)
0.38
(TRC)
0.35
(TRC)
0.35
(TRC)
1.18
0.153-
0.21
(FRC)
0.7
(TRC)
0.72
(TRC)
0.80
(CRC)
1.72
(FRC)
0.538
(CRC)
2.37
(CRC)
1.82
(CRC)
1.50
(FRC)
73,000
(TRC)
0.80
(FRC)
0.269
(FRC)
0.193
(CRC)
0.724
Test
Type
FT
FT
FT
FT
FT
S
FT
S/FT
S
FT
FT
.
-
-
S
S
FT
FT
FT
Kef arctic*
ZlOEeman, P.W., anc
Berg, R.O., 1934
Tc*i, C., and
McKee, J.A., 1978
Tsal, C.. end
McKee, J.A., 1978
Marking, L.L., and
Bills, T.P., 1977
Ward, R.W., and
De-Grave, G.M.. 1978
Ebellng, C., 1931
Brook, A.J.,.and
Baker, A.L., 1972
Markinp, L.L., and
Bills. T.D., 1977
Heath, A.C., 1977
Heath. A.C., 1977
Brooke, A.S., and
Seegert, G.L., 1978
Brooks, A.S., and
Seegert, C.L., 1978
Brooks, A.S., and
Seegtrt, G.L., 1978
Levlr, W.K., and
Ulrlch, H.C., 1967
Panlktar. B.M., 196C
Heath, A.G., 1977
Heath. A.C., 1977
Heatr.. A.G., 1«77
158
-------
Species
Colden Shiner
Colden Shiner
Colden Shiner
Colden Shiner
Colden Shiner
Colden Shiner
tt ii
ti ti
tt ti
n
n
tt ti
ti tt
ti »
ti ti
ti ii
ii ii
•i ti
•i ii
ii ti
ii it
•t ii
ii ii
•ndpoint
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
tl
tl
tl
n
Avoidance *
tt
ti
ti
ii
ti
ti
ti
if
ii
..
•I
(Mln.)
Exposure
Tine
5.760
5,760
1,800
810
8,640
8.640
2.880
2,880
8.640
8,640
10
II
II
It
II
II
tt
••
11
••
ii
it
T*op.
°C
24°
25°
5°
24°
f\
5°
24°
5°
24°
5°
24°
12°
18«>
24°
30°
12°
18°
24°
30°
12°
18°
24 *
30°
(Type)
Cone.
•g/L
(CRC)
0.930
(TRC)
0.04
(TRC)
0.84
(TRC)
0.26
0.18
0.18
(MCA)
0.99
(MCA)
1.09
(MCA)
0.64
(MCA)
0.92
(TRC)
0.209
(TRC)
MB
0.395
(TRC)
0.203
(CRC)
0.086
(CRC)
0.112
(CRC)
0.255
(CRC)
0.114
(FRC)
0.123
(FRC)
0.086
(FRC)
0.139
(FP.C)
0.088
Test
Type
FT
FT
_
V-.
•
-
-
-
_
V
-
FT
FT
FT
FT
FT
FT
FT
FT
FT
FT
FT
FT
Reference
Heath, A.G.. 1977
Heath, A.G., 1977.
Reath. A.C., 1978
Heath, A.C.. 1978
Heath, A.C., 197S
« ••- «i «
•i •" n
•t ti t*
ii n n
tl tl M
Larrirk, S.R., and
Cherry, T.S.,
and
Dickson, K.L., and
Cairns, J., 197S
•i ti »
ti ti ti
it ii n
ii ti ti
ii ii ti
ii ii ti
ii ii ti
ii ii ti
ii it it
Reproduced From
best available copy.
159
-------
Species
Golden Shiner
M n
ti fi
•t ti
ti »•
Coiranon Shiner
ti ti
ti tt
ti ti
•* «
Roseyfacc Shiner
ti ii
Spoctall Shiner
Tugnose Shiner
Spocfln Shiner
Emerald Shiner
•t ti
•i »i
ii 11
ii ••
Eadpolnt
Avoidance
«t
ti
tt
ti
LC-100
LC-50
LC-50
LC-50
LC-50
LC-100
LC-100
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
Ottn.)
Exposure
Tine
10
tl
n
M
12-30
76
5,760
2.880
2. 880
2,880
180
79
30
5,760
2.880
2.880
2,880
:,sso
"0
30
*r-
12°
18°
24°
30°
~*
250
10°
200
30°
—
—
10.1 -
20o
25°
10°
200
30°
10°
20°
30°
10°
25°
tType)
Cooc.
«g/L
(HOC1)
0.044
(HOC1)
0.027
(HOC1)
0.045
(HOC1)
0.033
(FRO
.015-
.017
(TRC)
0.7
0.51
(TRC)
0.78
(TRC)
0.59
(TRC)
0.45
(TRC)
0.07
(TRC)
0.07
Teet
Tjrp«
FT
FT
FT
FT
FT
FT
S
FT
FT
FT
FT
FT
FT
(TRC) S
2.41-0.53
(TRC)
fl.045
(TRC)
0.65
0.59
0.41
(TRC)
0.63
(TRC)
0.51
(TRC)
0.35
0.95
0.28
1
FT
FT
FT
tl
FT
FT
FT
FT
n
Reference
• 1 t. X
Larrick, S.R., aod
Cherry. D.S., and
Dlckson, K.L., and
Cains. J., 197b
n nn
• f- M W
Hubbs, L.L.. (1930)
Ward, R.K., and
Degrave, C.Y.. 1578
Brooks, A.S., and
Seegert. C.L.. 1978
tl 1* «t
tl II *t
Hubbs. C.L., 1930
ti *t n
Seegert, C.L., end
Brooks, A.S., 1978
Ward, R.V., end
DeCrave, C.K., 1978
Brooks, A.S., end
Seegert. C.L.. 1978
t* n it
Brooks. A.S., end
Seegert. C.L., 1978
•* tl •!
ii i. n
Fandrei, C.L., 1977
Fandrei. C.L., 1977
I Reproduced from S«*S
| best available copy. ^M^
160
-------
•pcelea
«•
V
Blunt Dose Minnow
(larvae)
Fathead Minnow
" (larvae)
Fathead Minnow
ft
"
n
ii
if
H
II
If
•1
II
ft
ft
Piraephsles Vigilar
Minnow
Rudd
Endpoint
LC-100
LC-60
DT.-68
510-
D. Repro.
LC-50
Ml
LC-50
LC-50
D. Repro.
LC-50
LC-50
D. Repro.
No Spawning
LC-100
LC-100
Reduced
Offspring
Avoidance
LC-100
LC-100
(Min.) Temp.
Exposure °C
Tine
61 —
43,200
43,200 —
10,080 —
5,760 —
5,760 —
7,200 —
720 —
100,800 —
5,760 25°
5,760 12°
_ —
__ __
__ —
4,320
10,080 —
—
79
2,460
(Type)
Cone.
•E/L
(TRC)
0.7
(CRC)
0.108
(CRC)
0.10S
(CRC)
0.043
(TRC)
0.08-0.19
(TRC)
0.05
(TRC)
0.02
(TRC)
0.185
(TRC)
0.110
(TRC)
0.082-.095
(TRC)
0.998
(CRC)
0.085
(CRC)
0.043
(CRC)
0.16-0.21
(CRC)
0.154
(CRC)
0.43
0.50
(TRC)
0.7
(FRC)
0.7
Test
Type
FT
FT
FT
FT
FT
FT
FT
FT
FT
FT
S
FT
FT
FT
FT
FT
~
FT
FT
Reference
Hubbs, C.L.,
1930
Arthur, J.V., and
Eaton, J.C.,
ii fi
II W
ZillicT., J.A
,« _•.*•
Basch, R.I..
1971
Authur. J.W.
1974
ft «t
1971
••
M
., 1969
M
et al.
, et al.
ti
Vard, R.K., and
DeGrave, C.K
Harking, L.L
Bills, T.B.,
Zillict. J.A
t.
tt tt
Arthur, J.W.
Eaton, J.G.,
it tt
Bogardus. R.
1978
Hubbs, C.L.,
Ebeling, C.,
., 1978
., and
1977
., 1969
tl
tl
, and
1971
11
B., et al
1930
1931
Reproduced from
besl available copy.
161
-------
Special
T«nch
Black Bullhead
Black Bullhead
* **
Oiannrl Catfish
« n
tt ti
« ti
tt «
n ti
ti n
ti ti
ti ii
tt ti
ft ii
Snail Mouth Bass
large Mouth Bass
"
Striped Bass
"
White Bass
n ii
Green Sirvflsh
Indpoiot
LC-20
LC-50
MT
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
Avoidance
LC-50
(Mm.)
Exposure
Ti»e
6.000
1.440
25
5.760
5.760
5.760
5,760
2.880
2.880
7.200
7.200
5.760
5,760
8,640
8.640
900
5.760
1,440
1.44Q
2.8SO
60
45
1.440
of '
—
120
6°
24°
12°
5°
2*>
5°
24°
6°
24°
6°
24°
—
25°
—
—
—
—
—
(Type)
Cone.
«g/L
(FRC)
0.7
(FRC)
^4.5
(TRL)
1.36
(TRC)
1.41
(FRC)
0.082
(FRC)
0.064
(TRC)
0.156
(FRC)
0.20
(FRC)
0.14
(FRC)
0.05
(FRC)
0.05
0.28
0.33
0.21
0.25
(FRC)
0.5
(TRC)
0.241
(TRC)
0.494
0.3
0.25
0. 25-. 035
0.0*5
(FRC)
2.0
Teat
Type
FT
S
FT
FT
FT
FT
S
FT
FT
FT
FT
FT
FT
FT
FT
S
FT
FT
S
S
FT
rr
s
Reference
Ebellng. C., 1931
Panikkar, B.M.. 1960
Truchan, J.G., and
Baach, R.E.. 1971
Marking. L.L
Bills, T.D.,
Heath, A.C.,
ti ti
Marking. L.L
Heath, A.G..
ti ti
" "
ii ti
tl •!
tl II
II tl
•1 II
Pyle. E.A..
Ward, R.W.,
DeGrave. C.K
Authur. J.W.
Hughes, J.S.
ti ••
Brungs, V.A.
Grieve, J.A.
1978
Panikkar. B.
.. end
1977
1977
n
., nod
(1978)
n
H
•i
•i
w
ii
I960
and
.. 1978
. 1974
. 1970
• 1
. (1973)
, et al .
V.. 1960
162
-------
Sped**
Green Sunfish
Sunfish
Green Sunfish
Grapple
Mosquito Fish
Eel
Black Grapple
Tellow Perch
Tellov Perch
tllow Perch
Walleye
ff
Minnows "Killies"
Notropis Volulellus
" fielnnis
Alewife
tf
ti
ti
ti
White Sucker
Cutthroat Trout
" (Juvenile)
Lake Trout
Endpolnt
LC-100
LC-50
LC-50
LC-50
MT
MT
MT
MT
LC-50
LC-50
LC-50
LC-50
NE
Avoidance
ti
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
•1
(Kin.)
Exoosure
Tine
5,760
5.760
5,760
4.320
6,000
25
65
720
5,760
720
5,760
120
—
. —
30
30
30
30
30
5.760
5,760
tf
TCMP.
°c
_
25°
12°
25°
—
_
—
—
^^
ISP
—
25°
—
—
—
10.5°
15.2°
20.1°
o
25.1
29.8°
120
10.1-
15.1°
12°
(Type)
Cone.
(CRC)
0.4
(TRC)
0.195-.278
(TRC)
1.2*
(TRC)
0.127
(FRC)
0.5-1.0
(TRC)
0.7
(TRC)
1.36
0.72
(TRC)
(TRC)
0.558
(TRC)
0.267
(TRC)
0.108
(FRC)
0.3
0.005
0.150
(TFC)
2.15
(TRC)
2.27
1.70
0.96
0.30
0.379
(TRC)
74.5-94.7
(TUC)
0.200
Test
Type
S
FT
5
FT
—
FT
FT
FT
FT
S
FT
FT
S
^_
—
S
S
S
S
S
S
FT
S
Reference
Coventry, F.L., et al
1935
Ward. R.V., and
DeCrave. O.K., 1976
Marking, L.L. , and
Bills, T.D., 1977
Ward, R.V., and
DeCrave, C.H., 1978
Cromou, A.S., 19*4
Ebeling, G., 1931
Truchan, J.G., and
Basch, R.E.. 1971
ft •! ,1
Authur, J.V., et al .
1974
Marking, L.L., and
Bills, T.D., 1977
Authur, J.W., et al.
1974
Vard. R.W., and
DeCrave, G.M. , 1978
Schaut, G.C., 1939
Bogr.rdus, R.B., et al
1978
It •! tl
Seegert, G.L., ar.d
BrookE, A.S., 1978
it it ,i
•i 11 11
ii •• M
Harking, L.L., and
Bills, T.D., 1977
Larsan, G.I.. , et al .
1978
Marking. L.L., ar.d
Bills. T.D.. 1977
Reproduced from ^py^
best available copy. ^UP
163
-------
•p*cic« Endpoint
1
Lake Trout LC-50
Blueglll LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
" LC-50
LC-50
" LC-50
Blacknose Dance Avoidance
Ottn.)
Exposure
Tine
5.760
5.760
5.760
5,760
2.880
2.880
2.880
2.880
10,080
_
5"*-
cc
l*t>
15-32°
6-32°
12°
10°
20°
30°
25°
32°
6°
25°
,__
(TypO
Cone.
(TRC)
0.060
(TSC)
-O.44
(TRC)
0.43-0.47
(T?.C)
0.555
(CRC)
3.00
(CFC)
1.72
(CRC)
1.23
(TRC)
0.54
(TRC)
0.47
(TRC)
0.33
(TRC)
0.37
(me)
Te«t
Type
FT
FT
FT
S
FT
FT
FT
FT
FT
TT
FT
^_
Reference
Ward. R.W.
DcGrave , C
Bans. M.L.
1977
Ei»». v.L.
Heath. A.C
Karklng, L
Bills, T.D
Brooks. A.
Seegtrt. C
•i
ti
Heath. A.C
ti
,i
•i
Fava, Teal
, and
.K., 1978
. et al.
. «tu«
.. 1*77
.L.. and
., 1977
S., and
.L.. 1978
II M
It ••
., 1978
•1 M
II II
II tl
1976
0.07.0.21.0.47
« ft
—
—
0
(MCA)
.07 and 0.17
—
•
M
ti tl
164
-------
APPENDIX C
SUMMARY OF TOXIC EFFECTS OF CHLORINE TO MARINE AQUATIC LIFE
>ecie6
Plants
Fhytoplankton
Chlorophyta (Algea)
Blue-green "
Tellow-green Algae
Astcrionella laoonia
n »
Chaetoceros declolens
** didvnium
betonula Confervace~a
Skeletonena costatu
ti ii
halassloslra Norden
" Pseudonona
it n
n ii
" Rotula
Monochrysls lutheri
Rhodomonas ballica
Phyto plankton
Giant Kelp
Invertebrate Ar.lmals
Sea anemone
riyCrold
Polychaete worm
Endpolnt
DG-71Z
DP-502
DC
DG-50
DG-50
DG-50
DG-50
DG-50
DG-50
DG-50
DG-50
DG-50
DG-50
DG-50
DG-50
DG-50
DG-50
DP-952
DP-502
NE
DG
(Mln.) Te«t
Exnosure Temp .
Time
240 —
_ _
5-10 —
1,440 -—
0.27 —
2 —
1,440 —
1,440 —
0.6
1,440
1.70 —
1,440 —
1,440 —
6.8
0.3 —
1,440 —
1,440
1,440 —
10
5,760
21,600
180
17" Decreased 5 —
(Type)
Cone.
ng/L
_
0.1
_
0.03
_
1.5
_
0.11
0.4
0.2
0.14
0.125
0.8
0.095
0.6
0.195
0.077
0.2
0.5
0.33
*
0.2
0.11
1.0
5-10
1.0
2.5
0.2
Test Reference
Type
— Carpenter, E.J., et •!,
1972
— Morgan, R.P. and
StroBB, R.G., 1969
— Hirayama, K., and
Hirano, R., 1970
— Gentile, J.H., et al.
1974
— Gentile, J.H., 1972
— ti it «
— Gentile, J.K., et al.
— 1974
— Gentile, J.R., 1972
— Gentile, J.H., et al.
1974
Gentile. J.H., 1972
— Gentile, J.H. et al.
_ 1974
Morgan, R.P. and
— Stress, R.G., 1969
ft it ti
Gentile, J.H., 1972
Gentile, J.H., et al.
1974
"~™ ' It tl «
*._ ** fl fl
Davis, M.H., and
Coughlan, J., 1978
— Clendenning, K.A., and
North, W.J., 1959
Turner, H.J., et al .
1948
McLean, R.I., 1972
— Huchmore, D. . r:id
T- i T» i m i
Sperm rortnlity
702 Decreaspd 5
Sperm mortality
0.4
165
-------
-»p«cie«
Crasnoctnea vir.
Oyster
CrassoKtnes Tlr.
Oyster
« M
Ostrea edul.
Oyster Larvae
Ostrea edul. Oyster
Mytllus edulis
(Mussel)
Mytllus edulis
(Mussel)
Mytllus edulis
(Mussel)
Castro pods
Aeartla ton SB
(Copepod)
Aeartla tons*
(Copepod)
« ii
•i
Eurvtenora aff.
(Copepod)
Eurvtemora aff.
(Copepod)
Pseudodlaptomus corn.
(Copeood)
PseudodlsDtomus coro.
(Copepod)
Adult copepod
n
II •!
Cooepod Stages
.i
.,
jcvarus aircicr.
(Decopcpod)
Endpoint
Ottn.) T.st
Exposure Tcs^i .
TIM
46~ Decreased — —
Ciliary
Punplng
Threshold
501 Decrease
In Tine open
Swimming
Stopped
HE
LC-100
LC-100
Unattached
Young
Stop Growth
LC-30
LC-70
LC-50
LC-50
LC-90
LC-50
" LC-50
LC-50
LC-50
LC-87
LC-69
LC-22
LC-70
LC-77
LC-26
LC-50
20-90
4,320 —
2 —
10 30°
21,600 —
7,200 —
— —
— _
2 20°
2 25°
120
0.7 —
5
360
2 —
45
5
2.PSO —
2,880
2 , 880
2,880
2,880 --
2.8SO —
5,760
(Typ.) T.st
Cone. Tree
-g/L
0.2
1.0
_ —
0.18
_
0.5
2.5 —
1.0
2.5 -
6.02-0.05 —
6.2 -
0.75 —
0*Jf
• /•>
1.0
_ _
10. n
_ _-
2.5
—
1.0
1 A f\
1 (.' . 1)
2.5 -
10.0 —
71.0 —
6.8-1,0 —
0-0.25 —
6.8-1.0
1.0 —
0-0.25
(CRC) —
0.05
Refer sac.
Galtsoff. P.S.
Calstoff, J.H
1946
Patrick, P...
..1946
;.. .t ai.
and
McLean, R.. 1970
Waugh, C.D.,
1964
Hydro. Bio Study 1967
Turner, H.J.,
1948
•i ii
James, V.C.,
White, W.R.,
Dressel, D.H.
Gentile, J.H.
1974
•i ii
McLean, R.I.,
Gentile, J.H.
1974
ii n
ii ii
Davis, M.H.
Couphlcn, J.t
ii n
..
ii it
•I ii
Capptizzo, J.M
1976
.t *1. .
n
1967
1966
, 1971
..
, et al.
n
1976
, et al.
n
"
tl
1978
fl
"
M
„
. , et al.
Reproduced from
best Available
166
-------
p.*..
PalaemcmetJE pcgls
(Decopepod)
Barnacle larvae
Barnacles
Barnacles Nauplil
•i ii
n ii
it «
n n
Cammarus lie.
(Aaphipod)
Me lit a Nitide
(Anphlpod)
tt II
Corophium Sp.
(Aaphipod)
•phlpod Juvenile
Anphipod
Blue crab
•i ii
Shape Crab
Sand Shrimp
(larvae)
n ii
Sand Shrimp
Shrimp
Shrimp (Juvenile)
Coonstripe Shrimp
Grass Shrimp
Mysic Shrimp
Bupula Sp.
ii 11
Sea Urchin
hiuroid
Endpolnt
LC-50
LC-80
LC-90-100
MT
No growth
LC-12
LC-58
LC-62
LC-25
LC-50
MT
HE
LC-50
LC-50
LC-50
LC-50
LC-50
LC-55
LC-42
LC-50
LC-50
LC-50
LC-50
LC-98
LC-50
LC-100
LC-100
DR-94-100i
Dk-22
(Min.) Teat
Exposure Tenp .
Time
5 -"i
21,600 —
10 —
10
2,880
2,880 —
2,880 —
180 —
120 —
5 —
410 ~
5,760 —
5,760 —
1,140
5,760 —
5,760
10 —
5
700
5,760
5,760
5,760
180
5,760
2, BBC
1,440
5
5 —
(Type) Test
Cone. Types
•g/L
0.22
2.5 -
1.0 —
6.5 -
1.0 —
6-0.25 —
0.8-1.0 —
71.0
2.5 —
2.5 —
5.5 —
10.0
0.687 ~
6.145 —
10.0
0.10 —
1.418 —
10.0
5.0 —
5.15 —
0.090
6.134 —
6.17B —
2.5 -
o.je.2
2.5
10.0
0.125
5.40 —
Reference
Bellanca, M.A..and
Bailey, D.S., 1977
McLean, R.I., 1973
Turner, B.J., et al.
1948
Waugh, C.P., 1964
ti ti ti
Davle, M.H., and
Coughlan, J., 1978
tl tl M
tl tl tl
Meldrln, J.W., et al.
McLean, R.I., 1973
II tl M
Gentile. J.H., 1972
Thacher. T.O., 1978
Thacher, T.O., 1978
Patrick. R., and
McLean, R. . 1970
Thacher, T.O., 1978
Gentile, J.R., et ml.
1974
" " *'
Patrick, and Kclean.R
1970
Thacher, T.O., 1978
lhacher, T.O., 1978
Thacher. T.O., 1978
McLean, R.I.. 1973
Thacher, T.O.. 1978
Turner, H.J., et al .
1 OAR
1 7*40
Muchmorc, P., 'and
Epol, D., 1973
•1 »• tl
167
Reproduced from
best available copy.
-------
Specie*
Echiuroid
totrvllus Sp.
Holnula Sp.
Fish
Tel low tall Flounder
Plaice (Larvae)
n
*i
ft
Winter Flounder
ii ii
Winter Flounder
eggs
Pink Salmon
"»ink Salmon
fink Salmon
Pink Salmon
Coho salmon
Coho salmon
Chinook Salmon
Chinook Salmon
Chinook Salmon
Chinook Salmon
Chinook Salmon
Young Salmon
Chinook Salmon
Atlantic Silver. side
•i ii
it ii
'Inncic Sllverside
»eggiO
Endpoint
DR-100Z
LC-100
LC-100
LC-50
LC-50
LC-50
LC-50
MT
LC-50
LC-50
LC-0
LC-50
LC-50
LC-50
LC-50
LC-50
- LC-50
Distressed
MT
MT
LC-50
LC-50
MT
LC-50
LC-50
LC-50
LC-50
LC-50
(Min.) Test
Exposure Temp.
Tine
5 —
1.440
4.320
1,440 —
5.760
460 —
75 —
4.320 —
15 —
0.3 —
20 —
5,760 —
7.5 13.6
15 13.6
5,760 —
7,200
5,760 —
60 —
130 —
23 —
7.5 11.7
30 11.7
33,123 —
5,7(0
90
30
5.760
2,880 —
frype)
Cone.
•g/L
6.40
10.0
1.0
0.1
(-)
0.028
(-)
0.05
(-)
0.075
0.25
2.5
10.0
16.0
6.05
6.5
6.25
(TRO)
n.023-.
0.08
(TRO)
0.032
0.10
0.25
1.0
0.5
0.25
0.05
0.038-
0.6.5
0.58
1.20
0.037
0.30
Test
Types
—
—
—
_ _
—
_
_
_
—
—
. .
—
—
__
052
—
_
—
—
—
»
—
- -
—
—
—
—
—
Reference
Muchmore, P.. and
Epel, D., 1973
Turner, H.J. , et •!.
1948
Gentile, J.H., et al.
1974
Alderson, R.. 1970
n n ti
•i ii n
ii ii ti
Gentile, J.H., et al.
1974
II II M
II II II
Holland, E.A., et al.
1960
Stober, Q.J., and
Hanson, C.H., 1974
Thacher, T.O., 1978
Holland, E.A., et al.
1960
Thacher, T.O., 1978
Holland, E.A.. et al.
1960
** II fl
•• •• It
Stober, Q.J., and
Hanson. C.H., 1974
Holland. I. A., et al.
1960
Thacher, T.O.. 1978
EngstroD, D.C., and
Kirkvood, J.B., 1974
Bellanca. M.A., and
Bally, D.S., 1977
Morgan, 7!. P., and
Prince. R.D., 1977
168
-------
Species
Tidewater Silverside
Blueback herring
Blueback herring
Blueback herring
(eggs)
Blueback herring
(Larvae)
Atlantic Menhaden
it ii
ii 11
ii ti
(Larvae) Atlantic
Menhaden
Threespine Stickleback
ii ii
White Catfish
'k>lden Shiner
flounder
Striped Mullet
(Juveniles)
Miscellaneous Marine
Fish
White Perch
White Perch
(eggs)
White Perch
(larvae)
Shiner Perch
English Sole
Pacific Sand Dance
Pacific Herring
Striped Sea Bass
Endpolnt
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50 _
LC-0
LC-50
LC-50
LC-50
LC-50
MT
MT
Irritant
Response
Death
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
(Mln.) Test
Exposure Temp .
Time
2,880 —
60
15 —
4,800 —
2 , 880 —
60 —
10
300 —
30 —
^ ••
5,760 —
5,760 —
2,880 —
5,760 —
e ^^
e ^^
-------
APPENDIX D
SUMMARY OF CHLORINE REACTION PRODUCT EFFECTS
ON FRESHWATER AND MARINE ORGANISMS
Specie*
(Invertebrate*)
Daphnli Baina
MM 11
Mldgel
Tanvtarsuc
diisHnilH
««...,
(Vertebrates)
lalnbou trout
•* M .1
Fathead ninnou
t. n ft
" " "
.uegill
it it
•i ii
(Plants)
Alga, Selena strum
capricornutum
H ii . it M
Endpolnt
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
LC-50
ELS
ELS
LC-50
LC-50
LC-50
EC-50
chlorophyll a
EC-50
cell number
Expo lure
Ti»e
48 hr.
48 hr.
48 hr.
48 hr.
96 hr.
96 hr.
96 hr.
96 hr.
96 hr.
96 hr.
96 hr.
96 hr.
96 hr.
96 hr.
96 hr.
96 hr.
Temp . Teat
CC Tvpe
- S, U
S, U
— S, U
— S, U
— FT, M
— FT, M
FT. M
— FT, M
— FT. M
S, U
S, U
— S, U
—
—
—
Compound
1.2
I.*
1.2
I.4
1,2
1,4
1.*
1.2
I,4
1.2
1.2
1.*
1,?
I.4
1.2
I,4
- 4KB
- PCB
- PCB
- >CB
- PCB
- PCB
- PCB
- *a
- PCB
- PCB
- PCB
-*CB
- »CB
- PCB
- PCB
- *CB
-g/L
Cone .
2.44
11.0
11.760
13.0
1.58
1.12
4.0
1.6-2.5
0.56-1.04
27.0
5.59
4.28
91.6
98.1
98.0
96.7
Reference
EPA 1978
EPA 1978
EPA 1978
EPA 1978
EPA 1980
EPA 1980
EPA 1980
EPA 1980
EPA 1980
Davson 1977
EPA 1978
EPA 1978
EPA 1978
EPA 1978
EPA 1978
EPA 1978
ELS • embryo larval stage
f. • measured
S ' static
I' • unmeasured
FT • flow through
170
-------
Species
(Invertebrates)
Mysld shrimp
Hysidopsis bahla
W If M It
Polychaetewonn
Polvdorm webster
Her Irs sp.
P. webster
P. Nereis
Tidewater
Silverside
Henidia beryllina
Sheepshead minnow
*!yprinodon
variegatus
Clan nercenaria
mercenaria
(embryo)
(larval)
Alga,
Skeletanema
costatum -
ti ii it
it ii M
Exposure Teap. Test
End point Time C Type
LC-50
LC-50
65Z emergence
from oysters
701
55Z
100Z
LC-50
LC-50
LC-50
LC-50
LC-50
EC-50
chlorophvll a
EC-50
cell number
96 hr. — S, U
96 hr. — S, U
3 hr. -- —
3 hr. -- —
3 hr. — —
3 hr. — —
96 hr. -- S, U
96 hr. — S, U
96 hr. — S. U
48 hr.
12 day _ _
96 hr.
96 hr.
Compound
1,2
1.*
1.2
1.2
1.*
1.4
1.2
1.2
1.4
1.2
1,2
1,2
l.A
1,2
1 ,4
- »CB
- PCS
- JCB
- »CB
- »CB
-*CB
- »CB
- >CB
- »CB
- DCB
- CCB
- *CB
- *CB
- PCS
- *CB
•g/L
Cone.
1.97
1.94
100.00
100.00
100.00
100.00
7.30
7.66
7.40
MOO.
>100
44.2
54.8
44.1
59.1
Kef erence
EPA 1978
EPA 1978
Mackenzie,
Shearer 1959
„ „ (1 I, M
1, „ ,( *• M
Dawson
EPA 1978
EPA 1978
Davis, Hindu
1969
EPA 1978
EPA 1978
EPA 1978
EPA 197E
171
-------
Species
End point
Exposure
Tine
Cone.
•«/*•
Test
Chemical
Reference
Plants
Algs,
Skeletonc
lostatun
Skeletonena lostatun
Skeletonena lo«t«tun
Skeletonena lostatun
Skeletonena loststun
Skeletonena lostatun
Skeletonema cosr.atua
Skeletoneaa costatun
Skeletonem coses tun
Skeletoneas coststun
Thalassloiiru
pseudonunu
Thalasslosiru
pseudonunu
Thalassiosiru
pseudonunu
Thalasslosiru
pseudonunu
Isoehrysis galbanu
Isochrvsis galbanu
Isochrvsis galbanu
Isochrvsis galbanu
Chlorophyll
a LC-50
fc-50
cell count
EC-50
cell count
F.C-50
cell count
EC-50
Chloro s
EC-50
cell count
25Z cell
inhibition
50Z cell
inhibition
25Z cell
inhibition
5CZ cell
inhibition
25Z cell
inhibition
50Z cell
inhibition
25Z cell
inhibition
50Z cell
inhibition
25Z cell
inhibition
50= cell
inhibition
251 cell
inhibition
50: cell
inhibition
96 hr.
96 hr.
96 hr.
96 hr.
96 hr.
96 hr.
7 days
7 days
7 days
7 days
7 days
7 days
7 days
7 days
7 days
7 days
7 days
7 davs
3.270
3.560
0.890
0.960
0.440
0.550
>8.0
>8.0
8.0
8.0
>8.0
>8.0
2.0
4.0
>8.0
>8.0
0.25
0.50
4, chlorc
i, chlorc
2,4,5 tri
2,4,5 tri
2,3,5.6 tetra
4 chloro
4 chloro
2,4,6 tri
2,4,6 tri
4 chloro
4 chloro
2,4,6 tri
2.4,6 tri
4 chloro
2,4,6 tri
2.4,6 tri
2,4,6 tri
— EPA :978
EPA 1978
— EPA 1978
— EPA 1978
— EPA 1978
S, U Erickson
S, U Erickson
S, U Erickson
S. U Erickson
S, U Erickson
S, U Erickson
S, U Erickson
S, U Erickson
S, U Erickson
S, U Erickson
S, U Erickson
S. U Erickson
Reproduced from
172
-------
APPENDIX E
SUMMARY OF TOXIC EFFECTS OF OZONE ON AQUATIC ORGANISMS
Residual
Species Ozone (mg/L) Summary of Effects Reference
Bluntnose Minnow
•—
Fathead minnow
Fathead ainnow
r Larvae
" Eggs
Fathead Minnow
•f it
ft It
M It
" "
larvae and eggs
larvae, egg' fish
adult fish
Central Common Shiner
••
"
it
Northern Common Shiner
tt it it
Pugnose Shiner
Spottsil shiner
post larvae
Rosyface Shiner
ti
Western Blacicnose
Dance
Korthem Creek Chub
it
" 0.03
0.09
0.25
0.50
0.50 ~~
1.25
0.2-0.3
<0.1
1.10
0.058
O.A7
0.225
0.059
0-0.016
0-0.013
0.016-0.01
0.001-0.003
0.03
0.25
1.25
1.84
0.003
0.016
0.016
1.22
0.03
1.25
O.C3
1.25
0.03
1.25
Slight IOBS of equilibrium
Death (56 Bin. exposure)
Loss of equilibrium (25 nin) ,
then death at unknown tine
Loss of equilibrium (10 Bin.)
Death, tine unknown
Nor death (20 nin.),
experiment stopped
100Z Death (1-3 hours)
-------
Residual
Cpeclcs 0,0,4. (mBn.1 Summarr of Effect* **f«r«tie«
Rainbow Trout
•i ••««
SB
Juvenile
•• M
•• «l
lalnbou- Trout
" " larvae
Lake Trout
ii
«
•i
"
"
Brown Trout
Cold Fish
•• it
" 11
•• ii
Splake (Salvellnus pp)
Chinook Salmon
H II
Coho Salmon
Bluegill
" larvae
ti ii
Bluegill
*'
Large Mouth Bass
White Sucker
egg
Yellow Perch larvae
«88
" " larvae
" " «86
White Perch
ii ii
•'
0.01-0.06
' 0.10
100! Mortality (4 hrs.)
Increased hatchability
0.0093 96 hrs.
-------
Species
Residual
Ocone ("g/L) SUBBUTV of Effect*
Reference
Channel Catfish
" larvae
" «gg»
Invertebrates
Cray Fish
Unspecified Flanktonlc
and Bottom Invertebrate
Daphnla Mayna
Stone Flies
Caddisflies
Aaphipod
Canaarus sp
laopods
Ascellus sp.
Snail
Coniobasis llvescers
Copepod
Cladoceran
Slnocephalus serrulatus
Marine I Esturiank
Aaerlcanoyster
eggs
Barnacles
Balanus sp
Phytoplawktou
Skeletonema costatum
Chlorella sp.
Nannochloris lutheri
Crab Zoea
Crab megalpos
Atlantic Silversidc
Striped Bass
eg?s
0.47
4.0
1.16
1.25
0.030
0.005
0.009
0.009
0.009
0.005
0.001
.001
<0.20
0.4-1.0
0.10
•
0.08
0.2
0.08-0.2
0.05-0.1
24 hr.
------- |