United States Office of Water November 1982
Environmental Protection Program Operations (WH 546) 430/9-82-010
Agency Washington DC 20460
&EPA Design of 301 (h)
Monitoring Programs
for Municipal
Wastewater Discharges
to Marine Waters
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DESIGN OF 301(h) MONITORING PROGRAMS FOR
MUNICIPAL WASTEWATER DISCHARGES TO MARINE WATERS
November, 1982
by
Tetra Tech, Inc., Staff
Contract Number 68-01-5906
Project Officer
Paul Pan, Ph.D.
Environmental Protection Agency
Washington, D.C. 20460
Tetra Tech, Inc.
1900 - 116th Avenue, N.E.
Bellevue, Washington 98004
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EPA REVIEW NOTICE
This report was prepared under the direction of the Office of Marine
Discharge Evaluation (WH-546), Office of Water Program Operations, Office of
Water, U.S. Environmental Protection Agency, 401 M Street, S.W., Washington
D.C., 20460, (202) 755-9231.
This report has been reviewed by the Office of Water and the Office of
Research and Development, U.S. Environmental Protection Agency, and approved
for publication. Mention of trade names or commercial products does not
constitute endorsement or recommendation for use.
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TABLE OF CONTENTS
LIST OF FIGURES v
LIST OF TABLES v1
CHAPTER I 1
OVERVIEW OF MONITORING REQUIREMENTS 1
INTRODUCTION 1
OBJECTIVES OF MONITORING 3
301(h) Requirements 4
State Requirements 5
NPDES Requirements 5
Other Requirements 7
CHAPTER II 8
TREATMENT PLANT AND EFFLUENT MONITORING 8
OBJECTIVES 8
SPECIFICATIONS FOR TREATMENT PLANT AND EFFLUENT MONITORING 9
Sampling Locations 9
Parameters 10
Sample Collection and Frequency 10
Analytical Methods for Toxic Substances 12
Toxic Substance Data Reporting 12
CHAPTER III 22
RECEIVING WATER QUALITY AND SEDIMENT MONITORING 22
OBJECTIVES 22
SPECIFICATIONS FOR WATER AND SEDIMENT MONITORING 22
Station Locations 22
Variables and Sampling Frequencies 25
Sampling and Analytical Methods 28
Oceanographic Measurements 30
Data Analysis and Reporting 34
CHAPTER IV 36
BIOLOGICAL MONITORING 36
OBJECTIVES 36
APPROACH AND RATIONALE 37
SPECIFICATIONS FOR BIOLOGICAL MONITORING 39
Sampling of Biological Communities 39
Station Locations 41
Sampling Frequency and Replication 45
Sample Collection and Processing 46
Analytical Techniques 75
Data Reporting 90
111
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CHAPTER V 92
QUALITY CONTROL 92
APPROACH AND RATIONALE 92
FIELD ACTIVITIES 93
QUANTITATIVE ERROR ANALYSIS 97
TOXIC POLLUTANT ANALYSIS 108
APPENDIX A 110
OCEANOGRAPHIC METHODS 110
CURRENT METERS 110
Uses of Current Meters 110
Types of Current Meters 111
DROGUES AND DRIFTERS 113
Use of Drogues 113
Types of Drogues 114
Uses of Surface Drifters 115
Types of Surface Drifters 116
Use of Seabed Drifters 117
Types of Seabed Drifters 118
DYE STUDIES 118
SPECIFICATIONS FOR FIELD WORK 119
Current Meters 119
Drogues and Drifters 120
Dye Studies 120
Navigation-Position Determination 122
REFERENCES 124
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LIST OF FIGURES
Number
1 Representative sampling locations for two levels of
biological monitoring 44
2 Examples of graphical displays of biological data from
a marine sewage discharge site 78
3 Oceanographlc surface observations log sheet 96
4 Example Precision Control Chart 100
5 Example Accuracy Control Chart 102
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LIST OF TABLES
Number Page
1 Monitoring Required by California Ocean Plan 6
2 Containers, Preservation, and Holding Times for Selected
Groups of Toxic Chemicals 13
3 List of Approved Analytical Methods for Selected Toxic
Chemicals (Priority Pollutants) 15
4 Example of Station Location Descriptions for 301(h)
Compliance Monitoring 26
5 Recommended Sample Preservation and Storage Requirements
for Water Quality 29
6 Recommended Analytical Methods 31
7 Selection of Appropriate Fish Sampling Gear 55
8 Examples of Some Nonparametric Statistical Tests 81
9 A List of Commonly-Used Indices of Diversity 86
10 A List of Some Common Polychaetes That Have Been
Associated with Marine or Estuarine Pollution 89
11 Recommended Procedures for Determination of Systematic
Bias and Precision in Analytical Methods 98
12 Summary of Expressions Necessary to Construct Control
Charts 104
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CHAPTER I
OVERVIEW OF MONITORING REQUIREMENTS
INTRODUCTION
Under Section 301(h) of the Clean Water Act of 1977 as amended by the
Municipal Wastewater Treatment Construction Grant Amendments of 1981,
publicly owned treatment works (POTWs) may apply for a variance from the
secondary treatment requirements for discharge into marine waters. Each
applicant is required to submit a detailed technical evaluation of the
discharge and its effects on the marine environment to demonstrate
compliance with the seven statutory criteria listed under Section 301(h).
If a variance were granted, monitoring would be required [Section 301(h)(3)]
to assess the impact of the modified discharge on marine biota. EPA
regulations implementing Section 301(h) are set forth in 40 CFR Part 125,
Subpart G, as amended in November, 1982.
The guidance provided in this document has been developed to help meet
the general monitoring requirements of the 301(h) program. References to
applicable water quality standards and requirements are not intended to
replace specific state requirements. Applicants must also check with the
appropriate state and local agencies for any specific monitoring
requirements applicable to their circumstances.
This document was prepared in order to provide guidance for designing
monitoring programs that will meet regulatory requirements (40 CFR 125.62)
and allow continuing assessment of the impact of less-than-secondary
discharges on the receiving water marine environment. It provides
supplemental guidance on designing monitoring programs to that included in
the Revised Section 301(h) Technical Support Document (Tetra Tech 1982)
which is available by writing to the Office of Marine Discharge Evaluation
(WH-546), U.S. Environmental Protection Agency, 401 M Street, S.W.,
Washington, D.C., 20460. The guidance provided in these documents is
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advisory only; its use is not required. However, EPA believes that Section
301(h) applicants will benefit substantially by following the guidance and
procedures provided in these documents.
The amended 301(h) regulations include a number of changes to the
monitoring program requirements contained in the original 301(h) regulations
promulgated in 1979. While the basic objectives of the overall monitoring
requirements remain the same, many of the original detailed requirements
were deleted from the amended regulations so that each applicant will have
the flexibility to design a cost-effective monitoring program to meet its
individual circumstances. This is especially true for small applicants that
discharge into depths greater than 10 meters with negligible seabed
accumulation of suspended solids [40 CFR 125.62(b)(2)].
Much of the guidance in this document is directed towards large
dischargers. It covers a wide range of possibilities that might be
encountered when developing 301 (h) monitoring programs, including complex
waste streams and discharges into sensitive ecosystems.
Users of this guidance document should keep in mind that the level of
effort for each 301(h) monitoring program must be keyed to the individual
circumstances of each discharge and corresponding receiving water situation.
A monitoring program will not have to be as extensive for smaller
dischargers as for large dischargers. A monitoring program for a waste
discharge comprised primarily of domestic wastes does not have to be as
comprehensive as a program to monitor the impact of a discharge with large
amounts of industrial and/or toxic wastes. The frequency of sampling
required for a resilient, high energy, or otherwise nonsensitive receiving
water environment will be considerably less than for a sensitive ecosystem.
A minimally acceptable monitoring program, then, will be based on a balance
of several factors, including the size of the discharge, the character of
the waste, and the sensitivity and variability of the receiving water
environment. In addition, a test of monitoring program practicability
should include consideration of the technical feasibility of available
measurement procedures during a variety of weather and sea conditions.
Those EPA tentative decision documents which recommend Section 301(h)
variances will highlight site-specific items which must be addressed in the
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applicant's proposed monitoring program. The 301(h) decision document
should, therefore, be analyzed carefully and the monitoring requirements
therein reflected in the design of the applicant's final monitoring program.
The Technical Evaluation or Technical Review Reports on individual 301(h)
applications should also be used as a reference in the design of final
monitoring program proposals.
OBJECTIVES OF MONITORING
Monitoring programs under 40 CFR 125.62 for dischargers receiving
modified NPDES permits under Section 301(h) of the Clean Water Act should be
designed to:
• Document short- and long-term effects of the discharge on
receiving water, sediments, and biota; also, on beneficial
uses of the receiving water
t Determine compliance with NPDES permit terms and conditions
• Assess the effectiveness of toxic control programs.
While divided into general biological, water quality, and effluent
monitoring components, in general, the monitoring program should focus upon
demonstrating the discharge's compliance with applicable standards and
permit conditions, and demonstrating predictable relationships between
discharge characteristics and impacts upon the marine receiving water
quality and the marine biota. Although each general monitoring component
may involve sampling at different locations for different variables and at
different times, it should not be considered as a separate and individual
activity, but as an integrated study. In this manner, the permittee should
be able to gain the most meaningful data on an assessment of the impacts of
the discharge. Further, once an adequate background data base is
established and predictable relationships among the biological, water
quality, and effluent monitoring variables are demonstrated, it should be
possible for many 301(h) permittees, especially those with small discharges,
to scale down the intensity of certain elements of their field monitoring
studies.
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Applicants may wish to expand their monitoring programs beyond the
minimum required to further demonstrate the impact or lack of impact of
their discharge on the environment. In addition, they may wish to exploit
unique opportunities provided by 301(h) to add to the body of knowledge on
effects of marine discharges on the receiving water environment and marine
ecosystems. The potential benefits to the municipality would be in
assessing long-range wastewater treatment and disposal needs and
alternatives. Additionally, applicants discharging in the same geographic
proximity may wish to develop an areawide assessment of marine discharge
environmental impacts. Applicants should consider, also, that the
monitoring data provided will be used by EPA to assess whether 301(h)
variances should be renewed following expiration of the initial
vari ances/permi ts.
301(h) Requirements
The monitoring requirements specified by 40 CFR Part 125.62 provide for
monitoring programs comprised of three elements: (1) biological monitoring,
(2) water quality monitoring, and (3) effluent monitoring. In addition,
applicants must demonstrate in their monitoring program proposals that they
possess the economic, personnel, technical, and other resources necessary to
implement their proposed programs. The biological and water quality
sampling must be able to detect variations over time and space as those
changes relate to the permittee's discharge. Monitoring must be conducted
at the current discharge site before and after any improvements are
implemented and at the site of new or relocated discharges. Sampling times
should include critical environmental periods and both typical and unusual
meteorological or oceanographic conditions. Biological programs for large
permittees and some small permittees must include field surveys of affected
or potentially affected biota, bioaccumulation studies, and an assessment of
the condition and productivity of commercial and recreational fisheries.
Water quality samples must be from stations selected to assess compliance
with water quality standards in the vicinity of the zone of initial dilution
(ZID), and beyond the ZID. The toxics monitoring program must determine the
effectiveness of industrial pretreatment and nonindustrial toxics control
programs. An adequate toxics monitoring program will also aid the
implementation of toxics control programs and the biological monitoring
efforts.
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The applicants must provide EPA with sufficient information on quality
assurance3 and control procedures to document compliance with accepted
scientific practice. The monitoring plan, therefore, should discuss quality
assurance in general, and specific data generation sections of the plan
should reflect individual details of quality control.
State Requirements
The monitoring program must document compliance with all applicable
water quality standards. Some states have specific monitoring requirements
and/or recommendations on parameters to be sampled, station locations,
sampling frequencies, and analytical methods. Table 1, for example, shows
the parameters recommended by the California Ocean Plan guidelines (CSWRCB
1972 and 1978). The parameters included in most State standards for
receiving waters are dissolved oxygen, pH, coliform bacteria, and suspended
solids or a surrogate. The standards have been expressed as a maximum
allowable pollutant concentration, a maximum allowable deviation from
background concentrations, a statistically significant difference between
stations, or as a prescription against harm to biota or degradation of
beneficial uses of the water body.
NPDES Requirements
An NPDES permit is required, by Section 402 of the Clean Water Act, for
all discharges of pollutants to navigable waters. The permits specify
effluent limitations plus effluent sample types (e.g., grab or 24-hour
composite) and sampling frequencies for assessing compliance. In some cases
the permits include receiving water monitoring requirements. The NPDES
requirements should be used by 301(h) applicants as a basis of decision
making on influent and effluent monitoring. NPDES permit sampling
a EPA policy initiated by the Administrator in a memorandum, dated May 30,
1979, stipulates that all environmental monitoring and measurement efforts
mandated or supported by EPA must have quality assurance project plans (see
Chapter V, Quality Control).
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TABLE 1. MONITORING REQUIRED BY CALIFORNIA OCEAN PLAN
Location of Monltorinq
Parameter
Flow
Bacteriological
Grease and Oil
Floatinq Participates
Suspended Solids
Settleable Solids
Turbidity
pH
Arsenic
Cadmium
Total Chromium
Copper
Lead
Mercury
Nickel
Silver
Zinc
Cyanide
Phenolic Compounds
Total Chlorine Residual
Ammonia Nitrogen
Total Identifiable
Chlorinated Hydrocarbon
Toxicity
Radioactivity
Salinity
Temperature
Biochemical Oxygen Demand
Total Phosphate
Total Nitrogen
Dissolved Oxygen
Discoloration
Light Transmittance
Water
Supply
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
Untreated
Wastewater
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
Effluent
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
X
Receiving
Water
X
X
x
X
X
X
X
X
X
X
X
Sediments
X
X
X
X
x
x
x
x
X
x
x
X
X
x
Fish and Macroinvertebrates
Sediment Sulphides
Particle Size Distribution
Benthic Biota
NOTE: X means monitoring required.
Source: California State Water Resources Control Board (1972, 1978).
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specifications will be supplemented (usually more frequent monitoring or
addition of parameters) to meet 301(h) objectives and/or state requirements.
Additional requirements to meet 301(h) objectives will consider 301(h)
related effluent limitations, plant flow characteristics, initial dilution
ratios, receiving water characteristics, and biological communities and
beneficial uses to be protected.
Other Requirements
For each discharge, an investigation should determine if there are
other water quality standards applicable to the given water body or other
monitoring program requirements. For example, the Interstate Sanitation
Commission has requirements relating to wastewater discharges in the New
York Harbor area. In California, basin plans developed by regional boards
or agencies contain requirements in addition to those found in the
California Ocean Plan.
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CHAPTER II
TREATMENT PLANT AND EFFLUENT MONITORING
OBJECTIVES
Plant monitoring (influent and effluent) is primarily required to
determine compliance with NPDES permit conditions and water quality
standards. In addition, influent and effluent monitoring provides
indicators for assessment of treatment plant performance. High effluent
pollutant concentrations may be due to plant malfunctions or overloads;
thus, plant monitoring can be used to identify problems and improve
performance. Effluent monitoring also provides information on waste
characteristics and flows for use in interpreting water quality and
biological data.
Monitoring programs for toxic substances and pesticides are required as
part of the 301{h) regulations and should be designed to:
• Determine the potential for toxicity to aquatic life and
risk to human health from toxic chemical substances
discharged to marine waters, and to
• Evaluate the effectiveness of industrial source control
pretreatment programs and nonindustrial toxic control
programs.
The first objective can be attained by measuring toxic chemical substances
1n the effluent and in selected samples taken from the receiving water
sediments and organisms used in biomonitoring protocols. The second
objective can be attained by measuring toxic substances in the treatment
plant influent and comparing trends. Sources of toxicants, whether
industrial or nonindustrial, are analyzed and identified as part of the
toxics control program.
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SPECIFICATIONS FOR TREATMENT PLANT AND EFFLUENT MONITORING
Major considerations in the design of plant sampling programs include
the specification of: sampling locations, parameters to be measured,
collection and analytical methods, and sampling frequencies.
Sampling Locations
Specific locations at the treatment plant for influent and effluent
sampling may be included in the NPDES permit. In the 301(h) monitoring
program only influent and effluent sampling points are specified, although
sampling at intermediate points within the plant may be useful for
monitoring individual treatment unit performance. Sampling at various
points in the collection system may also be necessary to isolate sources of
toxic substances.
For conventional pollutants and nutrients, influent samples should
generally be collected just downstream of the coarse screens or grit
chamber. If multiple waste streams enter the plant and a representative
sample cannot be collected, a flow-composite sample may be used for influent
analysis. Effluent samples should be collected downstream of any
chlorination or disinfection units. Samples should be taken as close to the
start of the outfall as possible. An example of such a sampling point would
be the effluent pumping station. Separate samples should be taken if two
outfalls are used and the effluent which enters the outfalls comes from
different parts of the treatment plant. When emergency bypasses are made to
a different outfall or discharge point, due to high inflows or treatment
plant problems, separate samples of the bypassed flows should be taken.
Sampling for toxic pollutants should include hourly grab samples
collected over a 24-hour period and composited in proportion to the flow.
Influent samples should be taken upstream of the plant intake works (prior
to the grit chamber, if possible) and the total (unfiltered) sample should
be analyzed. Effluents should be sampled after treatment and just prior to
entering the outfall pipe. If the effluent is chlorinated, samples should
be taken upstream and downstream of the chlorination unit.
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Parameters
The treatment plant monitoring parameters required by an NPDES permit
for a typical large discharger Include conventional pollutants, nutrients,
and toxicants. Influent monitoring normally includes volumetric flow rate,
BOD5, suspended solids, pH, and grease and oil. The suspended solids and
BODjj measurements are used to determine removal efficiencies and to detect
changes in the character of the waste stream. Measurements of pH and grease
and oil are used to determine the need for and success of any pretreatment
programs. Other influent monitoring parameters which may be required are
total phosphorus, total nitrogen or specific forms of nitrogen, settleable
solids, COD, and temperature. These variables may be needed to further
characterize the influent and to monitor treatment plant performance.
The parameters to be measured in the effluent include requirements of
the water quality standards, NPDES permit, and any additional variables
needed to interpret water quality and biological data. Plant effluent
monitoring should normally include volumetric flow rate, dissolved oxygen,
BOD5, suspended solids, settleable solids, temperature, total and fecal
coliform bacteria, grease and oil, and pH. If the effluent is chlorinated,
total chlorine residual is typically monitored. To aid in evaluating the
Importance of chlorination in forming persistent, possibly hazardous,
chloro-organics, a careful record of the total mass of chlorine used per
unit of flow should be reported. If other forms of disinfection are used,
type and dosage should be reported. Other variables which may be required
include floating particulates, total phosphorus, total nitrogen, ammonia or
other forms of nitrogen, and COD.
In addition to the above parameters, Section 125.62(d) of the amended
301 (h) regulations requires each applicant, to the extent practicable, to
monitor toxic substances and pesticides [see 40 CFR 125.8(u) and (m)] in the
effluent. A list of the 129 toxic pollutants is provided later in Table 3.
Sample Collection and Frequency
The type of sampling equipment to be used and sampling frequencies
depend on the size and nature of the discharge. Sampling frequency and type
of sample should be determined based on the variability of the influent and
10
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effluent characteristics and the flow so that the data collected will be
representative of the discharge. In general, volumetric flow rates of the
influent and effluent should be measured continuously using automatic
equipment. Hourly and average daily flow rates should be recorded. Daily
effluent and influent samples for BOD and suspended solids should be taken.
Twenty-four hour flow-composite samples are recommended. Nutrient sampling
may be done weekly or monthly using grab samples selected randomly or
24-hour flow-composite samples collected on randomly selected days during
the sampling period. Measurements of effluent pH should be done on daily
grab samples taken at different times each day. Daily grab samples are
typically taken for total and fecal coliform bacteria.
Generally, a randomly selected date within a defined sampling period,
coordinated with other sampling (i.e., for conventional pollutants) during
wet and dry flow periods, should be chosen for influent and effluent
sampling for toxic pollutants. Flows caused by bypass events at the POTWs
should also be considered for sampling and analysis.
The sampling frequency for toxic pollutants depends on such factors as
the size and location of the discharge, the types and quantities of toxic
pollutants present, and the sensitivity and beneficial uses of the receiving
water marine environment. More frequent sampling should occur for POTWs
with large discharges, significant types and quantities of toxic pollutants,
and sensitive receiving waters. All large POTWs and those small POTWs that
cannot certify they have no known or suspected sources of toxic pollutants
or pesticides need to establish baseline analyses for toxic pollutants and
pesticides present in their current discharge (40 CFR Part 125.64). Toxic
substance monitoring is required of all 301(h) waiver recipients to help
verify the type and quantity of the compounds identified in the discharger's
301(h) application and to determine if significant changes occur over time.
It is recommended that at least annual representative wet and dry
weather 24-hour composite sampling and analyses be undertaken. More
frequent sampling and analyses may be required depending on the type of
substances found in the wastewater and discharge or the sensitivity of the
receiving waters.
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More frequent plant monitoring may be necessary for both conventional
and toxic pollutants during the first year of the program to obtain reliable
estimates of when maximum waste load periods occur and the magnitude of peak
concentrations. In some cases the relationship between concentrations of a
variable (e.g., BOD or suspended solids) and volumetric flow rates is not
well known. Sampling at times of minimum, average, and maximum hourly flow
rates on a monthly basis for the first year should help define
concentration-flow rate relationships and allow better interpretation of the
receiving water quality data.
Analytical Methods for Toxic Substances
Regulations have been proposed on allowable holding times and
analytical procedures for toxic substances [45 Fed. Reg. No. 231
p. 79318-79379 (November 28, 1980)]. The final regulations establishing
test procedures for the analysis of toxic substances have not yet been
published. Recommended holding times, container requirements, and
preservation methods are listed in Table 2. Recommended analytical methods
are shown in Table 3 for the priority pollutants. Analytical methods for
the six pesticides (methoxychlor, mi rex, guthion, malathion, parathion, and
demeton) can be found in Watts (1980) and U.S. EPA (1978).
For many dischargers it will be necessary to contract with outside
laboratories for the analytical work. The laboratories selected should be
state certified according to U.S. EPA approved procedures. Sampling,
holding and analysis procedures, and equipment should comply with state and
federally approved methods.
Toxic Substance Data Reporting
Quality assurance information should be transferred quarterly to EPA
and should include copies of quality control charts used in the laboratory
and the results of replicate, split, spiked, and blank sample analyses. The
laboratory results submitted should include the calibration standards used,
copies of the calibration curves used, and the frequency of calibration
runs.
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TABLE 2. CONTAINERS, PRESERVATION, AND HOLDING TIMES
FOR SELECTED GROUPS OF TOXIC CHEMICALS
Parameter Container3
Metals
Chromium VI P»G
Mercury P»G
Metals in Table 3 (except above) P,G
Asbestos p
Cyanide (total and amenable p,G
to chlorination)
Organic Compounds
Extractable (including phthalates, G, teflon-
nitrosamines, organochlorine lined cap
pesticides, PCBs, nitroaromatics,
isophorone, polynuclear aromatic
hydrocarbons, haloethers, chlori-
nated hydrocarbons and TCDD)
Extractables (phenols) G, teflon-
1 ined cap
Purqeables (halocarbons and G, teflon-
Preservation
Cool , 4°C
HN03 to pH > 2
0.05% K2Cr20?
HN03 to pH > 2
1 ml 2.71% HgCl2
Cool , 4°C
NaOH to pH > 12
0.008% Na2S203e
Cool, 4°C
0.008% Na2S203
Cool, 4°C
H?S04 to pH > 2
07008% Na2S203e
Cool , 4°C
Maximum
Holding Time
24 hours
28 days
6 months
5 days
14 days
7 days
(until extraction)
30 days
(after extraction)
7 days
(until extraction)
30 days
(after extraction)
14 days
and aromatics)
1ined septum
0.008%
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TABLE 2. (Continued)
Parameter
Purgeables (acrolein and
acrylonitrile)
Pesticides
Phenols
Container3
G, teflon-
1 ined septum
G, teflon-
lined cap
P,G
Preservation
Cool, 4°C
0.008% Na2$203
Cool, 4°C
0.008% Na2S203e
Cool, 4°C
H2S04 to pH > 2
Maximum
Holding Time
3 days
7 days
(until extraction)
30 days
(after extraction)
28 days
Polyethylene (P) or Glass (G).
Sample preservation should be performed immediately upon sample collection. For composite samples each
alnquot should be preserved at the time of collection. When use of an automatic sampler makeliT Impossible
Sm??^6^!^04' ^^ S3mPleS ^ be PreS6rVed by m3intainin9 at 4ttC until compositing and sample
C Samples should be analyzed as soon as possible after collection. The times listed are the maximum times
that samples may be held before analysis and still considered valid. Samples may be held for longer periods
-6 "
Some samples may not be stable for the maximum time period given in the table. A permittee or
sanple for a shorter "™ if knowud9e exists ?° show ^
Guidance applies to samples to be analyzed by GC, HPLC, or GC/MS for specific organic compounds.
p
Should only be used in the presence of residual chlorine.
NOTE: If preservative is unavailable for organic compounds, recommended holding time is .48 hours at 4°C.
Source: U.S. Environmental Protection Agency (1979a).
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TABLE 3. LIST OF APPROVED ANALYTICAL METHODS FOR SELECTED
TOXIC CHEMICALS (PRIORITY POLLUTANTS)
Parameter and Units
1. Acenaphthene, ug/1
2. Acrolein, ug/1
3. Acrylonitrile, ug/1
4. Benzene, ug/1
5. Benzidine, ug/1
6. Carbon tetrachloride (tetra
chloromethane), ug/1
Methods (EPA Method Number)
GC or HPLC (610), GC/MS (625)
GC or HPLC (603), GC/MS (624)
GC or HPLC (603), GC/MS (624)
GC (602), GC/MS (624)
HPLC (605), Oxidation-
col ormetric, GC/MS (625)
GC (601), GC/MS (624)
Chlorinated Benzenes (other
than dichlorobenzenes)
7. Chlorobenzene, ug/1
8. 1,2,4-trichlorobenzene, ug/1
9. Hexachlorobenzene, ug/1
GC (601), (602), GC/MS (624)
GC (612), GC/MS (625)
GC (612), GC/MS (625)
Chlorinated Ethanes
10. 1,2-dichloroethane, ug/1
11. 1,1,1, trichloroethane ug/1
12. Hexachloroethane, ug/1
13. 1,1-dichloroethane, ug/1
14. 1,1,2-trichloroethane, ug/1
15. 1,1,2,2-tetrachloroethane, ug/1
16. Chloroethane, ug/1
Chloroalkylethers (chloromethyl,
chloroethyl and mixed ethers)
17. Bis (chloromethyl) ether3
18. Bis (2-chloroethyl) ether, ug/1
19. 2-chloroethyl vinyl ether
(mixed),ug/1
GC (601),
GC (601),
GC (612),
GC (601),
GC (601),
GC (601),
GC (601),
GC/MS (624)
GC/MS (624)
GC/MS (625)
GC/MS (624)
GC/MS (624)
GC/MS (624)
GC/MS (624)
GC (611), GC/MS (625)
GC (601), GC/MS (624)
Chlorinated Naphthalene
20. 2-chloronaphthalene
GC (612), GC/MS (625)
15
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TABLE 3. (Continued)
Chlorinated Phenols (other than those
listed elsewhere, includes trichloro-
phenols and chlorinated cresols)
21. 2,4,6-trichlorophenol, ug/1
22. Para-chloro meta-cresol, ug/1
23. Chloroform (trichloromethane), ug/1
24. 2-chlorophenol, ug/1
GC (604), GC/MS (625)
GC (604 , GC/MS 625)
GC (601), GC/MS (624)
GC (604), GC/MS (625)
Dichlorobenzenes
25. 1,2-dichlorobenzene, ug/1
26. 1,3-dichlorobenzene, ug/1
27. 1,4-dichlorobenzene, ug/1
GC (601, 602, 612), GC/MS (625)
GC (601, 602, 612), GC/MS (625)
GC (601, 602, 612), GC/MS (625
Dichlorobenzidine
28. 3,3-dichlorobenzidine, ug/1
HPLC (605), GC/MS (625)
Dichloroethylenes
29. 1,1-dichloroethylene, ug/1
30. 1,2-trans-dichloroethylene, ug/1
31. 2,4-dichlorophenol, ug/1
GC (601), GC/MS (624)
GC (601), GC/MS 624)
GC (604), GC/MS (625
Dichloropropane and Dichloropropene
32. 1,2-dichloropropane, ug/1
33. 1,2-dichloropropylene (1,2-dichloro-
propene), ug/1
34. 2,4-dimethylphenol
GC (601), GC/MS (624)
GC (601), GC/MS (624)
GC (604), GC/MS (625)
Dinitrotoluenes
35. 2,4-dinitrotoluene, ug/1
36. 2,6-dinitrotoluene, ug/1
37. 1,2-diphenylhydrazine, ug/1
38. Ethyl benzene, ug/1
39. Fluoranthene, ug/1
GC (609), GC/MS (625)
GC (609), GC/MS (625)
GC/MS (625)
GC (602), GC/MS (624)
GC or HPLC (610), GC/MS (625)
16
-------
TABLE 3. (Continued)
Haloethers (other than those listed
elsewhere)
40. 4-chlorophenyl phenyl ether, ug/1 GC (611), GC/MS (625)
41. 4-bromophenyl phenyl ether, ug/1 GC (611), GC/MS (625)
42. Bis (2-chlorisopropyl) ether, ug/1 GC (611), GC/MS (625)
43. Bis (2-chloroethoxy) methane, ug/1 GC (611), GC/MS (625)
Halomethanes (other than those
listed elsewhere)
44. Methylene chloride (dichloromethane), ug/1 GC (601), GC/MS (624)
45. Methyl chloride (chloromethane), ug/1 GC (601), GC/MS (624)
46. Methyl bromide (bromomethane), ug/1 GC (601), GC/MS (624)
47. Bromoform (tribromomethane), ug/1 GC (601), GC/MS (624)
48. Dichlorobromomethane, ug/1 GC (601), GC/MS (624)
49. Trichlorofluoromethaneb
50. Dichlorodifluoromethaneb
51. Chlorodibromomethane, ug/1 GC (601), GC/MS (624)
52. Hexachlorobutadiene, ug/1 GC (612), GC/MS (625)
53. Hexachlorocyclopentadiene, ug/1 GC (612), GC/MS (625)
54. Isophorone, ug/1 GC (609), GC/MS (625)
55. Naphthalene, ug/1 GC or HPLC (610), GC/MS (625)
56. Nitrobenzene, ug/1 GC (609), GC/MS (625)
Nitrophenols
57. 2-nitrophenol, ug/1 GC (604), GC/MS (625)
58. 4-nitrophenol, ug/1 GC (604), GC/MS (625)
59. 2,4-dinitrophenol, ug/1 GC (604), GC/MS (625)
60. 4,6-dinitro-o-cresol, ug/1 GC (604), GC/MS (625)
Nitrosamines
61. N-nitrosodimethylamine, ug/1 GC (607), GC/MS (625)
62. N-nitrosodiphenylamine, ug/1 GC (607), GC/MS (625)
63. N-nitrosodi-n-propylamine, ug/1 GC (607), GC/MS (625)
64. Pentachlorophenol, ug/1 GC (604), GC/MS (625)
65. Phenol, ug/1 GC (604), GC/MS (625)
17
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TABLE 3. (Continued)
Phthalate Esters
66. Bis (2-ethylhexyl) phthalate,
67. Butyl benzyl phthalate, ug/1
68. Di-n-butyl phthalate, ug/1
69. Di-n-octyl phthalate, ug/1
70. Diethyl phthalate, ug/1
71. Dimethyl phthalate, ug/1
ug/1
GC (606)
GC (606)
GC (606
GC (606
GC (606
GC/MS (625)
GC/MS (625)
GC/MS (625)
GC/MS (625)
GC/MS (625)
GC (606), GC/MS (625)
Polynuclear Aromatic Hydrocarbons
72. Benzo (a) anthracene (1,2-benzy-
anthracene), ug/1
73. Benzo (a) pyrene (3,4-benzopyrene), ug/1
74. 3,4-benzofluoranthene, ug/1
75. Benzo (k) fluoranthene
(11,12-benzofluoranthene), ug/1
76. Chrysene, ug/1
77. Acenaphthylene, ug/1
78. Anthracene, ug/1
79. Benzo (ghi) perylene (1,12 benzo-
perylene), ug/1
80. Fluorene, ug/1
81. Phenanthrene, ug/1
82. Dibenzo (a,h) anthracene
(1,2,5,6-dibenzanthracene), ug/1
83. Indeno (1,2,3-cd) pyrene, ug/1
84. Pyrene, ug/1
85. Tetrachloroethylene
(tetrachloroethene), ug/1
86. Toluene, ug/1
87. Trichloroethylene
(trichloroethene), ug/1
88. Vinyl chloride (chloroethylene), ug/1
GC or HPLC (610), GC/MS (625)
GC or HPLC (610), GC/MS (625)
GC or HPLC (610), GC/MS (625)
GC or HPLC (610), GC/MS (625)
GC or HPLC (610), GC/MS (625)
GC or HPLC (610), GC/MS (625)
GC or HPLC (610), GC/MS (625)
GC or HPLC (610), GC/MS (625)
GC or HPLC (610), GC/MS (625)
GC or HPLC (610), GC/MS (625)
GC or HPLC (610), GC/MS (625)
GC or HPLC (610), GC/MS (625)
GC or HPLC (610), GC/MS (625)
GC (601), GC/MS (624)
GC (602), GC/MS (624)
GC (601), GC/MS (624)
GC (601), GC/MS (624)
Pesticides and Metabolites
89. Aldrin, ug/1
90. Dieldrin, ug/1
91. Chlordane (technical mixture
and metabolites), ug/1
GC (608), GC/MS (625)
GC (608), GC/MS (625)
GC (608), GC/MS (625)
18
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TABLE 3. (Continued)
DDT and Metabolites
92. 4,4'-DDT, ug/1
93. 4,4'-DDE (p,p-DDX), ug/1
94. 4,4'-ODD (p.p-TDE), ug/1
GC (608), 6C/MS (625)
GC (608), GC/MS (625)
GC (608), GC/MS (625)
Endosulfan and Metabolities
95. a-endosulfan-Alpha, ug/1
96. B-endosulfan-Beta, ug/1
97. Endosulfan sulfate, ug/1
GC (608), GC/MS (625)
GC (608), GC/MS (625)
GC (608), GC/MS (625)
Endrin and Metabolites
98. Endrin, ug/1
99. Endrin aldehyde, ug/1
GC (608), GC/MS (625)
GC (608), GC/MS (625)
Heptachlor and Metabolites
100. Heptachlor, ug/1
101. Heptachlor epoxide, ug/1
GC (608), GC/MS (625)
GC (608), GC/MS (625)
Hexachlorocyclohexane (all isomers)
102. a-BHC-Alpha, ug/1
103. B-BHC-Beta, ug/1
104. Y-BHC (lindane) -Gamma, ug/1
105. 6-BHC-Delta, ug/1
Polychlorinated Biphenyls
106.
107.
108.
109.
110.
111.
112.
PCB-1242 (Aroclor
PCB-1254 (Aroclor
PCB-1221 (Aroclor
PCB-1232 (Aroclor
PCB-1248 (Aroclor
PCB-1260 (Aroclor
PCB-1016 (Aroclor
1242), ug/1
1254), ug/1
1221), ug/1
1232 , ug/1
1248), ug/1
1260), ug/1
1016), ug/1
GC (608), GC/MS (625)
GC (608), GC/MS (625)
GC (608), GC/MS (625)
GC (608), GC/MS (625)
GC (608),
GC (608),
GC (608),
GC (608),
GC (608),
GC (608),
GC (608),
GC/MS (625)
GC/MS (625)
GC/MS (625)
GC/MS (625)
GC/MS (625)
GC/MS (625)
GC/MS (625)
19
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TABLE 3. (Continued)
Miscellaneous Substances (including
metals, organic compounds not listed
elsewhere, and asbestos)
113. Toxaphene, ug/1 GC (608), GC/MS (625)
114. Antimony (total), ug/1 AA (204.2)
115. Arsenic (total), ug/1 AA (206.2)
116. Asbestos (fibrous), chrysotile TEM
fibers MFL (million fibers per liter)
117. Beryllium (total), ug/1 AA (210.2)
118. Cadmium (total), ug/1 AA (213.1)
119. Chromium (total), ug/1 AA (218.1 or .2 or .3)
(IV), ug/1 AA (218.4)
120. Copper (total), ug/1 AA (220.2)
121. Cyanide (total), ug/1 Titrimetric, Spectro-
photometric (335.2)
122. Lead (total), ug/1 AA
123. Mercury (total), ug/1 AA
239.2)
245.1 or .2)
249.2)
124. Nickel (total), ug/1 AA
125. Selenium (total), ug/1 AA (270.2)
126. Silver (total),ug/1 AA (272.2)
127. Thallium (total), ug/1 AA (279.2)
128. Zinc (total), ug/1 AA (289.2)
129. 2,3,7,8-tetrachlorodibenzo- GC/MS (613), (625)
p-dioxin (TCCD), ug/1
a Bis (chloromethyl) ether was removed from the toxic pollutant list
(U.S. EPA 1981a).
b Dichlorodifluoromethane and trichlorofluoromethane were removed from the
toxic pollutant list (U.S. EPA 1981b).
Note: GC = Gas chrom tography.
HPLC = High performance liquid chromatography.
GC/MS = Gas chromatography coupled with mass spectrometry.
Source: U.S. EPA (1979a).
AA = Atomic absorption.
Source: U.S. EPA (1979b).
TEM = Transmission electron microscopy.
Source: Anderson and Long (1980).
20
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The results of screening measurements for the priority pollutants in
the effluent should be included in the monitoring reports. The results of
the chemical analyses should, where possible, be reported as measured rather
than less-than-certain values. If the analytical results were below the
limit of detection, this should be noted on the data sheet and the value
given as less than the actual limit of detection (e.g., < 10 ug/1). The
list of compounds identified in previous screenings should be compared to
the new results. The estimated concentration after initial dilution should
be computed for each toxicant. These values after initial dilution should
be compared to available criteria for marine waters [45 Fed. Reg. No. 231
pp. 79318-79379 (November 28, 1980)]. Those compounds which exceed the
criteria should be added to the list of toxicants subject to bioassay.
Toxicants found which can be bioaccumulated, but not previously included in
the sediment monitoring program, should be added to that program.
21
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CHAPTER III
RECEIVING WATER QUALITY AND SEDIMENT MONITORING
OBJECTIVES
To determine compliance with water quality standards and the 301(h)
criteria, the receiving water quality monitoring program must document water
quality in the vicinity of the Zone of Initial Dilution (ZID) boundary, at
control or reference stations, and at areas beyond the ZID where discharge
impacts might reasonably be expected. Monitoring must reflect conditions
during all critical environmental periods as identified in the 301{h)
applications. If currently available data are not adequate to predict when
critical periods will occur, then greater monitoring effort may be necessary
to demonstrate that water quality data are collected under the appropriately
critical conditions. Examples of such critical conditions are periods of
anadromous fish spawning runs, juvenile fish migrations or feedings, high
wastewater loadings, high water temperature, and low flushing rate.
The applicant's historical sampling programs, together with new
requirements associated with 301(h) permit conditions, will be the most
useful guide for designing an adequate receiving water monitoring program.
Changes in station locations, parameters, or frequencies may be required to
rectify deficiencies in historical programs.
SPECIFICATIONS FOR WATER AND SEDIMENT MONITORING
Station Locations
Section 125.61 of the amended 301(h) regulations requires that water
quality be maintained to assure the protection of public water supplies, the
protection and propagation of a balanced indigenous population (BIP) of
shellfish, fish, and wildlife, and to allow recreational activities. Under
Section 125.62, the establishment of a water quality monitoring program is
required which to the extent practicable:
22
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• Provides adequate data for evaluating compliance with
applicable water quality standards
• Measures the presence of toxic pollutants which have been
identified or are reasonably expected to be present in the
discharge.
In order to meet these water quality monitoring requirements, receiving
water and sediment sample stations need to be located in the vicinity of the
ZID boundary, at control sites, and 1n Impact areas in such a way as to
allow adequate correlations to be made between water quality, oceanographic
and sediment measurements, and toxic substances and biological data. Other
locations which a state may wish to specify include the shoreline in
swimming and shellfishery areas and within the ZID. Placing stations so
that pollutant concentration gradients can be detected between the ZID
boundary and control stations may be valuable for larger discharges.
When a discharge is into a saline estuary there is a greater emphasis
on protecting benthic organisms within the ZID, suggesting that water
quality data near the seabed and sediment quality data may be necessary. In
the case of oceanic discharges there are general requirements to prevent
extreme adverse biological impacts within the ZID which have adverse effects
beyond the ZID; thus, again some water quality monitoring within the ZID may
be required for especially sensitive ecosystems and/or large or
industrialized wastewater systems.
Criteria for selection of specific stations depend on the purpose of
the station. ZID-boundary stations should be placed on the upcurrent and
downcurrent boundaries of the ZID; they will not necessarily be at fixed
locations but more likely will be set on the day of sampling based on
observations of current direction. Since the objective is to intercept the
waste field drift flow as it is carried across the ZID boundary, several
samples placed at depth and across the wastefield need to be obtained. This
will be necessary in order to compute an average value and to show a range
of values if the waste field is not uniform. To demonstrate that the plume
has indeed been sampled, especially if this is not evident by the water
quality values themselves, data on currents, drogue tracks, and/or tracers
need to be provided.
23
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Stations upcurrent of the ZID may be the choice for water quality
control stations, although with the gradient concept in mind, stations
sufficiently far downcurrent may be satisfactory. Care should be taken in
selecting control stations so that values presumably representing control
conditions do not include diluted wastes carried back into the area of the
control station by tidal currents. Control stations should be unaffected by
other pollutant sources as well as the applicant's discharge. Also, control
stations should be located in water of similar depth as the discharge, with
similar bottom characteristics and similar distances from shore. Additional
controls may be needed when the applicant's proposal is for a relocated
outfall and when the discharge is into stressed waters.
Impact area stations vary from one discharge site to another. Areas
where monitoring may be required include recreational beaches, diving areas,
shellfish harvesting areas, kelp beds, coral reefs, commercial and
recreational fishing grounds, and other distinctive biological habitats.
The selection of impact area stations should be based upon a thorough review
of the recreation and biological sections in the 301 (h) application, the
Technical Evaluation or Review Report, the tentative decision document, and
the draft 301(h) permit. State requirements on station locations need to be
met by the program (e.g., both shore and nearshore stations must be sited to
protect beaches under the California Ocean Plan).
Additional stations may need to be located near other pollutant sources
to allow the effects of the subject discharge to be distinguished from these
sources. Examples of other pollutant sources are areas off the mouths of
major rivers near the discharge, sludge disposal areas, and other municipal
and industrial ocean discharges. The need for these stations is identified
by noting the extent of influence from available water quality data, from
analysis of potential impacts based on volumetric flow and characteristics
of the discharge, and from analysis of the dispersion characteristics of the
receiving water body.
All sited stations should be plotted on large-scale nautical charts or
15-min quadrangle sheets (USGS) and then transferred to more convenient
small scale maps. Latitude and longitude should be determined from the
large maps to the nearest second. The approximate depth at each station
24
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should be determined from previous sampling data or estimated from soundings
shown on the large-scale charts. Each station location should be described
relative to the outfall/diffuser, permanent navigation buoys, or distances
from known shoreline points as shown in Table 4. Historical designations
and/or the applicant's designation should also be noted. If ZID boundary
stations are occupied using accurate navigation methods there should be
adequate assurance that the resulting water quality sampling data reflect
ZID boundary conditions. To document station locations, the applicant's
program and periodic monitoring data reports should describe navigation
(locating) methods and field conditions during sample collection.
Variables and Sampling Frequencies
The variables to be sampled in the receiving water include those
specified in the 301(h) regulations, those required by the state, and those
necessary to evaluate other water quality data. The variables which should
be included routinely are BODg dissolved oxygen, pH, temperature, salinity,
suspended solids or its surrogates (e.g., light transmittance), total and
fecal coliform bacteria, and settleable solids. Light transmittance may be
specified in terms of turbidity, Secchi disc depth, extinction coefficient,
or percent light transmittance. The applicant should state the reason(s)
for the light transmittance method(s) selected. Additional variables which
may be required are:
• Total nitrogen
—nitrate
—nitrite
—total kjeldahl nitrogen
—ammonia
• Total phosphorus
—reactive phosphorus
t Chlorophyll £
• Floating particulates
• Color.
25
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TABLE 4. EXAMPLE OF STATION LOCATION DESCRIPTIONS
FOR 301(h) COMPLIANCE MONITORING
Station Historical Approximate
Number N Latitude W Longitude General Description Designation depth, roa
1
2
3
4
21°15'10"
21°17'44"
21°17'37"
21°17'46"
157°49'53"
157°52'34"
157°53'22"
157°53'59"
3.14 km directly south of Waikiki
Beach, approx. 2.4 km west-southwest
of Diamond Head Beach Park.
0.79 km directly south of Sand Island
in Honolulu Channel, alonq west edge.
at end of old sewer outfall, 1.07 km
offshore from Sand Island.
along eastern edge of Kalihi Channel,
None
2b
3b
4b
70
12.5
12.fi
12.fi
1.0 km directly southwest of coral reef,
0.7 km directly west of diffuser.
5 21°17'44" 157°54'25" 2.74 km directly southeast of Ahua Pt., 5b 8.5
0.66 km directly west pf Kalihi Channel,
0.60 km directly south of Keehi Lagoon
coral reef.
6 21°17'0r 157°54'24" at the center of the zone of nixing None 68
48 m north of diffuser within the ZID.
7 21°17'01" 157°53'59" ^0.7 km directly east of center of zone None 70
of mixing, at the eastern edge of ZID.
8 21°16'56" 157°54'24" 130 meters directly south of center of None 79
zone of mixing at ZID boundary.
9 21°17'01" 157°55'00" 1.024 km directly west of center of zone S1-9C 128
of mixing, just outside west edge of
zone of mixing.
10 21°17'16" 157°54'24" at north edge of zone of mixing, 0.48 km S1-6C 31
directly north of center of zone of
mixing, 0.53 km north of diffuser.
11 21°17'39" 157°54'57.5" 2.33 km south-southeast of Ahua Pt., None 15
1.56 km directly west of Kalihi Channel,
0.83 km directly south of Keehi Lagoon
coral reef.
12 21°17'08" 157°53'22" 0.91 km directly south of old sewer out- None 70
fall, 1.77 km east of center of zone of
mixing, 1.81 km directly south of coral
reef at south end of seaplane runway.
13 21°17'22" 157°55'18" 2.68 km south of Ahua Pt., 1.69 km west- None 70
northwest of center of zone of mixing,
1.54 km directly south of Keehi Lagoon
coral reef.
14 21°17'54" 157°55'46" 1.64 km directly south of Keehi Lagoon Hone IS
Beach, 2.86 km directly northwest of
center of zone of mixing.
15 21°17'27" 157°52'26" 1.19 km directly south of southernmost None 18
tip of Sand Island, approx. 500 m south-
southeast of entrance to Honolulu Channel.
16 21°15'H" 157°49'01" 0.76 km southeast of Diamond Head Beach lb 18
Park near lighthouse and Coast Guard Res.
17 21°17'01" 157°53'48" 1.0 km directly east of center of zone None 70
of mixing, just outside eastern edge of
zone of nixing.
a All depths are relative to MLLW.
b Applicant's proposed monitoring station numbers.
c Historical sampling stations of applicant.
26
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Water column profiles for salinity and temperature are needed to
interpret dissolved oxygen data and may be necessary to predict the location
of the drift flow unless profiling with tracers or effluent constituents is
to be relied upon for finding the plume. In any event, salinity and
temperature may be needed, together with current speed and direction, to aid
in describing mass movement of diluted wastes to farfield sites where
impacts must be assessed. Current directions are needed to determine where
in the horizontal plane one might expect to find the drift plume crossing
the ZID boundary. Current speeds are important in assessing the vertical
height of plume rise and, hence, in establishing where in the vertical plane
samples should be taken to measure plume constituents. Salinity,
temperature, and currents, furthermore, need to be provided to assist in
evaluating the results of benthic and other biological responses. In
estuaries, the amount of freshwater inflow from rivers needs to be
documented as an adjunct to evaluating residence times and routes of
possible transport of diluted effluents and particulates. Sampling may be
at generally accepted depths, e.g., 1 m below the water surface, mid-depth,
1 m above the sea bed and at 10-m intervals for depths greater than about 40
m; however, features of water masses observed in profiling for salinity and
temperature (and possibly light transmittance) should take precedence in
establishing sample depths.
Parameters to be measured in the sediments should include particle size
distribution and total volatile solids. Other variables, such as BOD5,
sulfides, and total organic carbon, may be required by the states or may be
important in analyzing discharge impacts on benthic biota. In addition,
sediment samples should be analyzed annually for toxic substances and
pesticides identified in the plant effluent. Sediment samples for toxic
substance analysis should be taken within the ZID, in the vicinity of the
ZID boundary, at representative impact area locations outside the ZID
boundary, and at control stations.
Sampling in estuaries should be conducted at slack water as recommended
in the Revised Section 301(h) Technical Support Document (Tetra Tech 1982).
Where tidal effects are to be discriminated, sampling should be done at
several times over a tidal cycle for both spring and neap tides. In order
to verify continuing compliance with 301(h) criteria, the ZID boundary
27
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stations should be sampled during those times of year when the discharge is
least diluted.
Sampling frequencies must be selected to meet state requirements and
should provide data during the critical environmental period(s) as
identified in the 301(h) application. For sites where available data do not
define these periods adequately, receiving water sampling should be done
monthly for at least the first year. In most other cases, quarterly surveys
which cover the critical period(s) should suffice. More frequent sampling
may be specified by the states in swimming and shellfishery areas to
determine compliance with bacteriological standards.
Sampling and Analytical Methods
The monitoring programs should specify sample collection, preservation,
storage, and analysis methods which are approved by the EPA and state
agencies and are appropriate for the site. In addition to specification of
analytical methods, the minimum accuracy, limits of detection, and desired
number of significant digits to be recorded should be specified to help
ensure that accurate and precise data are obtained. Receiving water samples
should be taken with a Van Dorn, Frauchy, or comparable sampler and then
transferred to the proper type of container. When variables are measured by
electronic probe (e.g., temperature, conductivity, dissolved oxygen, and pH)
values should be measured at 1- to 3-m (3- to 10-ft) intervals. Electronic
probe systems often have severe accuracy problems. Frequent calibration is
essential.
Sediment samples for organic carbon should include only the upper 2 cm
(0.8 in) of the sediment to ensure that the sediment oxygen demand per unit
mass is not diluted by underlying, stabilized, or inorganic material.
Table 5 lists sample preservation and storage requirements for some
variables, showing minimum sample volume, type of container, preservative
required, and maximum storage time. Any deviations in container type or
preservative from those specified in this table should be noted on sample
bottles, field data sheets, and in the monitoring reports. The volumes
given are intended as minimum amounts. Sample volumes should be increased
depending on the number of sample splits to be made.
28
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TABLE 5. RECOMMENDED SAMPLE PRESERVATION AND
STORAGE REQUIREMENTS FOR WATER QUALITY
Parameter
Oil and Grease
Total Suspended
Solids
Settleable Solids
Volume
Required (ml)
1.000
Sufficient aliquot to con-
tain residue of >_ 25 mg.
Successive aliquots of
sample may be added to
the same dish.
1,000
Extinction Coefficient 100
pH
Ainnonia Nitrogen
Salinity
Temperature
Biochemical Oxygen
Demand
Total Phosphorus
Reactive Phosphorus
Nitrate - N
Nitrite - N
25
400
240
1,000
1,000
50
50
100
50
Total Kjeldahl Nitrogen 50-500
Dissolved Oxygen
Color
Total and Fecal
Col i form
Total Organic Carbon
Chlorophyll a_
a If samples cannot
reported data should
300
50
100
25
200
be returned to the laboratory
Indicate the actual holding
Container
G only
G, P
P. G
P, G
P. G
P. G
G (with paraffined
corks)
P, G
P. G
P. G
P, G
P. G
P. G
P, G
G only
P. G
P, G, sterilized
P, G
P, not acid-washed,
keep in darkness
away from light
In less than 6 hours
time.
References: Standard Methods (American Public Health Association 1980)
Preservative
Holding Time
Analyze immediately or 24 hours
5 ml HC1 at time of „
collection. Cool, 4° C
None. Analyze
immediately.
None
Analyze same day or
cool, 4° C
Determine on site or
cool , 4° C ~
Cool . 4° C
H2S04 to pH <2
None
Det. on site
Cool . 4° C
Cool, 4° C
H2$04 to pH <2
Filter on site.
Cool , 4° C
Cool, 4° C
H2S04 to pH <2
Cool, 4° C
Cool, 4° C
H2S04 to pH <2
Determine on site
Cool , 4° C
Add Na thiosulfate
to effluent samples
Cool, 4° C
H2S04 to pH <2
Filter on site, add
MgC03 during
filtration
and holding time exceeds
and Methods for Chemical
No holding
24 hours
7 days
6 hours*
7 days
1 hour
(Longer If properly
sealed air tight)
No holding
24 hours
24 hours
24 hours
24 hours
>24 hours
24 hours
24 hours
No holding
24 hours
Analyze 1n field
or 6 hours
24 hours
Process Immedi-
ately or 2 weeks
if frozen and
kept in dark.
this limit, the final
Analysis of Water
and Hastes (U.S. EPA 1979b).
29
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Chemical analysis procedures for sediments are basically the same as
for water samples once the sediments have been digested. The EPA/Corps of
Engineers manual, Procedures for Handling and Chemical Analysis of Sediment
and Water Samples (Plumb 1981), should be consulted for detailed guidance on
sediment sample handling and digestion procedures. Sediment samples can be
stored dried, frozen, or on ice for metals analyses. If analysis of organic
constituents is to be performed, sediment samples should be stored on ice
only.
Analytical methods should be selected based on the water quality
standards and the EPA-approved methods listed in 40 CFR Part 136, with due
consideration of the extent to which interferences occur in the receiving
water and wastewater samples. When several methods are available, the
selection should be made by comparing the accuracy and precision of the
candidate methods for the concentration range expected at the site, and the
adequacy of the detection limit of each method relative to the pertinent
water quality standard. Table 6 shows recommended methods for the same
variables listed in Table 5 along with acceptable minimum precision and
detection limits for each method. The table also shows the desired number
of significant digits to be reported for each parameter.
Oceanographic Measurements
Oceanographic measurements to meet the 301(h) objectives include two
parts: data needed to detect plume and sediment movement, and observations
needed to interpret the water quality and biological data. This section
discusses acquisition of current, wind, and tide data. Verification of
initial dilution calculations is not required. However, if plume
calculations are considered by the applicant to be unreliable or inaccurate,
it may be desirable to obtain supplemental monitoring data to improve the
models or to document field validation of other models.
The POTW's 301(h) application and its Technical Evaluation or Review
Report should be reviewed to determine if available current data and
knowledge of wind and tide effects are adequate to determine the direction
of movement of the wastefield beyond the ZIO boundary, the subsequent
dilution, and the direction of movement of the sediment from the discharge.
30
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TABLE 6. RECOMMENDED ANALYTICAL METHODS
Parameter
Flow
Grease and Oil
Floating Participates
Total Suspended Solids
Settleable Solids
Extinction Coefficient
discoloration
oH
r ' '
Salinity
Temperature
Dissolved Oxygen
Biochemical Oxygen
Demand
Chemical Oxygen
Demand
Total Organic Carbon
Ammonia Nitrogen
Nitrate-Nitrogen0
Method
Continuous measurement, auto-
matic device
Gravimetric, Separatory Funnel
Extraction EPA Method 413.1
Flotation funnel extraction
Gravimetric, Dried at 103-105° C.
EPA Method 160.2 (SM, 14th ed..
p. 91, Section 206D)
Volumetric, Imhoff Cone.
EPA Method 160.5.
Gravimetric Method (SM 14th
ed. , pp. 95-96, Sec. 208F)
Light transmissometer
Presence or absence of color
at surface
Potentiometric. EPA Method
150.1 (SM 14th ed. , p. 460,
Sec. 424)
Induction Salinometer
or titration. (SM 14th ed.,
p. 107, Sec. 2G9C)
Bathythermograph or Thermo-
metric. EPA Method 170.1.
(Si: 14th ed., p. 125, Sec. 212)
Modified Winkler, Full Bottle
Technique, with azide modifica-
tion fur effluent sampler o_r
Membrane Electrode when cali-
brated with Modified Kinkier
with azide modification for
effluent samples. EPA Method
360.2 or EPA Method 360.1 when
calibrated with EPA Method 360.2
(SM 14th ed., pp. 441-447. Sec.
422 A and B for Winkler; SM 14th
ed., p. 450, Sec. 422F for probe)
5 day, 20° C
EPA Method 405.1
(SM 14th ed., p. 543, Sec. 507)
EPA Method 410.3
(SM 14th ed., p. 550,
Sec. 508)
Combustion-Infrared Method
EPA Method 236 (SM 14th ed.,
p. 532, Sec. 505)
Automated Phenate Method.
EPA Method 350.1 (SM 14th ed..
p. 616, Sec. 604 or Strickland
and Parsons, p. t>7)
Technician Auto Analyzer 11 px
Spectrophometric, manual,
Cadmium Reduction. EPA Method
3t>3.2 or EPA Method 353.3.
(SM 14th ed., p. 423, Sec.
419c, for manual method)
Significant
Minimum Digits
Precision Detection Desired
+ 8 percent 0.02 MGD
+_ 0.9 mg/1 5 mg/1
N/Aa mg/m2
N/A mg/m
approx. +_ 5 mg/1 10 mg/1
N/A 1 ml/l/h
N/A
N/A N/A
+0.1 standard unit pH = 12
titration: +_ 0.05 ppt 1 ppt
+_ 0.05° C
+ 0.05 mg/1 (Winkler) 0.1 mg/1
Probe: +0.1 mg/1 0.1 mg/1
+ 0.7 mg/1 BOD at 1 mg/1
2 mg/1 BOD
+ 26 mg/1 BOD at
175 mg/1 BOD
+ 13 mg/1 COD 1 mg/1
N/A N/A
+ 0.005 mg NHj-N/l 0.01 mg NHj-N/1
Automated, 0.05 mg/1
Cadmium:
+ 0.092 mg/1 at
0.35 mg/1
Spectrophometric,
Cadmium:
+ 0.004 mg/1 at
0.24 mg/1
3
2
2
2
2
2
N/A
3
4
3
3
2
2
2
2
2
31
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TABLE 6. (Continued)
Parameter
Nitrite-Nitrogen
Total Kjeldahl
Nitrogen
'Total Phosphorus
Method
Technician Auto Analyzer II or
Diazotization. EPA Method
304.1 (SM 14th ed.. p. 434.
Sec. 420)
Technician Auto Analyzer II or
Colorimetric, EPA Method 351.3.
(Sh 14th ed., p. 437. Sec. 421)
Technician Auto Analyzer II or
Digestion, Manual Ascorbic Acid
Technique. EPA Method 365.2
(SM 14th ed., p. 476, Sec.
42bC(III) and p. 481 Sec. 425F
for manual method)
Precision
N/A
Colorimetric
+ 1.056 at 4.10
mgN/1
Automated:
+ 0.130 mgP/1 at
0.8 mgP/1
Manual :
+ 0.033 mgP/1 at
0.11 ngP/1
Significant
Minimum Digits
Detection Desired
0.01 mg H02-N/1 2
Colorimetric 2
<1 mgN/1
0.005 mgP/1 2
Fecal Coliform
Total Coliform
Chlorophyll a_
Particle Size
Distribution"
Multiple Tube Fermentation
Technique, MPiJ Test. (SM 14th
ed., p. 922, Sec. 908C)
Multiple Tube Fermentation
Technique, MPH Test (SM 14th
ed., p. 916. Sec. 906A) or
for seawater only Membrane
Filter (SM 14th ed.. p. 928,
Sec. 9CW)
Spectrophotometric (SM 14th
ed., p. 1029, Sec. 1002G
Strickland and Parsons, SCOR/
UNESCO Equation, pp. 185-194)
Sieve Analysis (Buchanan)
MPN with 95 percent NA
confidence limit
HPN with 95 NA
percent confidence
limit
NA NA
8 N/A = not available.
Minimum accuracy, when given as a range about a specific value, has been taken from the below listed refer-
ences. The associated ranges are for 1 standard deviation about the mean value.
The cadmium reduction method determines nitrate + nitrite-nitrogen. The nitrate-nitrogen 1s calculated by
subtracting nitrite-nitrogen as determined by a separate diazotization test.
Detailed procedure is given in the biological monitoring section.
Reference's:
Buchanan, O.B. 1971. Sediments. In: International Biological Program (IBP) Handbook No 14
Blackwell Scientific Publ., Oxford, pp. 30-52.
CLMBS: Great Lakes Region Comittee on Analytical Methods. 1969. Chemistry Laboratory manual
bottom sediments. EPA, Federal Water Quality Administration. Washington, D.C., 101 pp.
EPA Method: U.S. Environmental Protection Agency. 1979b. Methods for Chemical Analysis of
Water and Wastes. USEPA, Environmental Support Laboratory, Cincinnati. OH.
SM 14th ed.: American Public Health Association. 1980. Standard methods for the examination of
water and waste water. 14th ed., Washington, D.C. 1193 pp.
Strickland and Parsons: Strickland, J.D.H.. and T.H. Parsons. 1972. A practical handbook of
seawater analysis. Dilution 167, 2nd ed. Fisheries Research Board of Canada. Ottawa, Canada.
310 pp.
32
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The current data should be reviewed to determine whether surface, bottom,
and mid-depth currents were measured. The sampling times should be reviewed
to determine whether periods of minimum and maximum stratification and other
important conditions (e.g., onshore winds, upwelling periods) were covered.
Winds generally are an important influence on coastal currents.
Current data should therefore be checked to see if historical surface
currents were measured concurrent with wind measurements and to what depth
wind effects are discernible. The wind data should be reviewed to estimate
frequency of occurrence of onshore transport by season and location. This
information is helpful in identifying discharge impact areas along the
shoreline. Any deficiencies in available current data should be noted and a
determination made as to whether intensive current monitoring is needed or
if specific data gaps need to be filled.
Field observation methods selected for oceanographic measurements
depend on the kinds of information needed, the extent of potential discharge
impacts, the oceanographic and physical conditions at the site, and
resources available to the applicant. Drogues, drifters, or dye released
from the outfall site may be used to determine mean current velocities at
specified depths, and also to provide information on the direction of
movement and the dispersive properties of the velocity field. These studies
may also identify nearshore eddy patterns or "dead" circulation zones which
may be present. Drogues set just above the bottom, or seabed drifters, can
be used to determine the direction of sediment movement.
Oceanographic data which should be recorded at the time of water
quality and biological sampling include:
t Wind speed and direction
t Sea state (height of swell and waves).
A variety of field study methods are available to collect oceanographic
information. The appropriate use and limitations of current meters, drogues
and drifters, dye studies, and field positioning methods are discussed in
Appendix A to this document.
33
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Data Analysis and Reporting
Data reporting procedures include the preparation of field logs, sample
container labels, laboratory data sheets, and reporting forms. The
reporting forms should be completed and sent to the appropriate U.S. EPA
Regional Office on the schedule prescribed in the 301(h) permit. Chain of
custody forms, showing the transfer of data from the field to the
laboratory, and finally to the U.S. EPA, should be maintained along with
field logs and laboratory data sheets.
The type of information recorded on the field logs and sample labels
should ensure that samples are identified properly and data are recorded
accurately. Field logs should include station location and number, depth of
samples, type(s) of samples taken, date and time of sampling, surface
observations as specified in the oceanographic section, depth of water at
the station where samples are taken, all field water quality measurements,
and the names of all individuals who collected the samples. This
information should be entered on the field log at the time of sampling. The
sample container labels should give sample number, station location, and
number; date, time, and depth of sample; treatment of sample (e.g., ^$64
added); a code designating what analyses are to be done on the sample, and
the name of the individual(s) collecting the sample. Laboratory data sheets
should include sample number, station location and number, sampling date and
time, and name of the analyst. For each individual analysis, results should
be reported along with the unit of measurement, duration of sample storage,
date sample was analyzed, and any comments on deviations from laboratory
procedures or unusual sample conditions. The chain of custody forms should
show the name of the person to whom the form is being sent; and the name of
the person receiving the data and date received.
Receiving water quality and sediment data should be compared with NPDES
requirements (when applicable) and applicable water quality standards.
Spatial gradients should be examined to determine whether elevated
concentrations occur near the outfall and, if so, where concentrations
return to background levels. Analysis of temporal trends should be done to
identify seasonal differences. Appropriate statistical tests (e.g., ANOVA)
can be used to determine if statistically significant differences exist
between the ZID-boundary and reference (control) stations. The water
34
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quality and sediment data can be used to define and report on the spatial
extent of the wastewater plume and sediment deposition area. This
information should be used in conjunction with the biological monitoring
data to identify and interpret any changes detected in the biota.
35
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CHAPTER IV
BIOLOGICAL MONITORING
OBJECTIVES
Biological monitoring is necessary to evaluate the overall impact of
the permittee's modified discharge. The primary objective of the biological
monitoring program is to provide evidence that:
• There is a continued attainment or maintenance of water
quality which assures protection and propagation of a
balanced, indigenous population (BIP) of shellfish, fish,
and wildlife beyond the zone of initial dilution (ZID) and
in the vicinity of the ZID boundary
• Conditions within the ZID do not contribute to extreme
biological impacts, such as the destruction of distinctive
habitats of limited distribution (e.g., kelp beds and coral
reefs), the presence of disease epicenters, or the
stimulation of phytoplankton blooms which have adverse
effects beyond the ZID, etc.)
t For discharges into saline estuarine waters: a) benthic
populations within the ZID do not differ substantially from
balanced, indigenous populations which exist in the vicinity
of the ZID boundary, b) the discharge does not interfere
with estuarine migratory pathways within the ZID, and c) the
discharge does not result in an accumulation of toxic
pollutants or pesticides at levels which exert adverse
effects on the biota within the ZID
t There is a continued attainment or maintenance of water
quality which allows for recreational activities (including
36
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fishing) beyond the ZID boundary and such activities in the
vicinity of the modified discharge are not restricted unless
such restrictions are routinely imposed around sewage
outfalls discharging secondary effluent.
Design of the biological monitoring program requires careful
consideration of potential impacts specific to the 301(h) permitee's
discharge(s). Factors which are important to designing biological
monitoring programs and individual sampling procedures are discussed below.
APPROACH AND RATIONALE
Section 125.62(b) of the amended 301 (h) regulations requires that the
biological monitoring programs for both small and large 301(h) discharges
must provide date adequate to evaluate the impact of the modified discharge
on the marine biota. This generally necessitates comparing the
characteristics of selected marine communities in the vicinity of the
discharge with the characteristics of similar communities in reference
areas. Therefore, the same type of comparative strategy required for
demonstrating a balanced, indigenous population (BIP) of shellfish, fish,
and wildlife in the application should be incorporated into the biological
monitoring program. [See the Revised Section 301(h) Technical Support
Document (Tetra Tech 1982) for guidance on demonstrating a BIP in a 301{h)
application.]
Under Section 125.62(b)(l)(i-iv) of the amended regulations, biological
monitoring programs must to the extent practicable include:
• Periodic surveys of the biological communities and
populations most likely affected by the discharge, as well
as those at suitable control sites, to enable comparisons
with baseline conditions
• Periodic determinations of the accumulation of toxic
pollutants and pesticides in organisms and examination of
adverse effects such as disease, growth abnormalities,
physiological stress, or death
37
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• Sampling of sediments in the vicinity of the ZID boundary,
in other areas of expected sediment accumulation, and at
appropriate reference sites to support water quality and
biological surveys and to measure the accumulation of toxic
pollutants and pesticides
• Periodic assessment of the conditions and productivity of
commercial or recreational fisheries where the discharge
would affect such fisheries.
Except for the periodic survey requirement, small permittees are not subject
to these specific requirements if they discharge at depths greater than 10 m
and can demonstrate through a suspended solids deposition analysis that
there will be negligible sea bed accumulation in the vicinity of the
modified discharage. However, small permittees still must provide adequate
data to evaluate the impact of the modified discharge on the marine biota.
This should involve the establishment of a background data base and the
demonstration of predicted biological impacts of the small discharge. In
all cases, site-specific characteristics will affect the selection of the
number of sampling sites, sampling locations, and the required sampling
effort in each biological category.
Information available in the discharger's 301(h) application and in
other investigations conducted near the discharge should be utilized fully
to identify physical-chemical and biological characteristics of the
potentially affected receiving waters. Characterization of the
oceanographic and meteorological setting of the discharge area will be
necessary to make decisions concerning positioning of the discharge and
reference sampling stations. Available biological data should be reviewed
to define limits of natural variability in biological populations. The
number of sampling stations and number of replicate samples at each station
should be determined, in part, on the basis of this information. In these
respects, the historical data may serve the same purposes as a pilot survey.
Decisions concerning taxonomic groups to be sampled, station locations,
types of sampling equipment, sample handling, sorting procedures, and
ancillary measurements of physical-chemical parameters should be made on the
basis of existing information to the extent practicable. Where effects of
the proposed discharge on specific biological communities or important
38
-------
species has not been clearly resolved, the monitoring program should be
designed to fill such data gaps.
In designing the program specifications, the complementary nature of
the water quality and biological monitoring programs should be recognized.
Concurrent collection of biological and water quality information should be
emphasized in an effort to identify causal relationships.
SPECIFICATIONS FOR BIOLOGICAL MONITORING
Sampling of Biological Communities
The 301(h) regulations require "periodic surveys of the biological
communities and populations which are most likely affected by the discharge
to enable comparisons with baseline conditions described in the
application." Emphasis should generally be placed on monitoring of benthic
communities due to the inherent community characteristics, sampling
considerations, and the importance of the benthos in the marine ecosystem.
Benthic communities adjacent to pollution sources can generally provide
information on the area! extent of impact more readily than other biological
communities because many benthic organisms are sedentary or relatively
immobile and are, therefore, continually exposed to pollution stress.
Benthic communities are also more easily sampled than other biological
communities and benthic sampling methods are more standardized than methods
for other communities. Existing information on benthic communities is often
sufficiently extensive to provide documentation of both the magnitude and
direction of the community response to specific perturbations. Finally,
benthic communities are of a primary importance in the food chain of the
nearshore marine environment. For the above reasons, monitoring of the
macrobenthos should normally be a primary element of 301(h) permit
biological monitoring programs.
Another principal monitoring requirement defined in the regulations is
periodic assessment of the accumulation of toxic pollutants and pesticides
in the biota. These assessments are required as part of a specific
monitoring effort for measuring the impact of elevated or increasing levels
of toxic pollutants and pesticides on susceptible biological communities.
Bivalves (e.g., My til us californianus and M. edulis), have been shown to
39
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greatly concentrate most identified marine pollutants relative to ambient
concentrations in seawater. Because water quality characteristics (e.g.,
temperature and dissolved oxygen) affect biological uptake, the
concentrations of toxic substances in the tissues of these organisms will
more accurately reflect the site-specific potential for bi©accumulation than
will the measurement of ambient concentrations of toxic substances. For
these reasons, caged bivalves used in offshore biomonitoring systems may
provide an early warning of excessive water column contamination, and may be
used to monitor the potential for transfer of toxic pollutants and
pesticides into and through the food chain. Such an in situ biomonitoring
system also provides a means of evaluating the effectiveness of toxic
control programs.
Monitoring program requirements also include the periodic assessment of
commercially or recreationally important fisheries that may be affected by
the discharge. The objective of fisheries monitoring is to assess the
condition and productivity of those fisheries. Fisheries monitoring may
consist of the periodic review of catch data collected by state agencies,
interviews of sport fishermen to determine success rates, or field and
market sampling of the fish or shellfish populations.
The biological monitoring program regulations specify that the
permittee monitor the biological communities and populations which are
likely to be affected by the discharge using comparisons with baseline
conditions described in the application. In addition to the benthos, such
communities may include phytoplankton, zooplankton, and fishes.
Numerous site-specific characteristics of the environment may
necessitate additional biological monitoring. For example, hydrographic
characteristics (current patterns, water residence time) and nutrient
concentrations in an estuary or embayment may result in the potential for
long-term biological changes such as eutrophication. If such changes are
determined to be a potential impact of the modified discharge, periodic
monitoring of the phytoplankton community may be required. Furthermore, if
changes in species composition of the phytoplankton occur and thereby induce
changes in the species composition of the herbivore community, zooplankton
monitoring may be required.
40
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In certain geographic settings, some coordination of the monitoring
programs of adjacent dischargers may be required so that those dischargers
periodically sample the same biotic groups. These conditions include, for
example, the situation where the potential for eutrophication exists, or
where further examination of the multiplicative effects of several
dischargers on widespread toxic dinoflagellate blooms is needed.
Monitoring should be initiated to ensure the continued existence of
distinctive habitats of limited distribution. Such habitats include, but
are not limited to, kelp beds, coral communities, and rocky intertidal
communities.
Station Locations
To meet the minimum requirements of the biological monitoring program,
sampling of the selected biological communities in the vicinity of the ZID,
in any other areas of expected impact, and at appropriate control sites will
generally be required. Other sampling locations which may be specified
include nearfield areas where important habitats have been identified, and
also at both new and old discharge sites in the case of an improved
discharge involving outfall relocation. In the case of large permittees,
additional sampling is recommended at intermediate locations between the ZID
boundary and control stations along a gradient of effluent concentrations to
help define the spatial extent of biological effects.
Information derived from the water quality monitoring program will be
important in interpreting the results of biological sampling. Therefore,
the selection of water quality and biological stations must not be done
independently. In addition, the station selection process of the
biomonitoring program should place emphasis on the inclusion of historical
sampling sites. This will maintain sampling continuity, and additional
information will thus be made available for impact assessment.
In instances where chemical analysis of the effluent and/or sediments
in the vicinity of the discharge has identified toxic pollutants at levels
of concern, the required bioaccumulation monitoring of toxic substances
should be undertaken in the vicinity of the ZID boundary, at other areas of
expected impact, and at control site(s). Test organisms utilized for in
41
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situ blomonitoring studies should be placed at the depth of the plume during
the exposure period. If the plume surfaces, exposure should be conducted at
a sufficient distance below the surface to prevent damage to or loss of the
exposure apparatus. Additional exposure depths may be used if the plume
depth is uncertain or variable, or if past discharges have resulted 1n
substantial sediment contamination. In cases of past sediment
contamination, near-bottom exposures can be used to evaluate the
contribution of sediments to bioaccumulation levels.
The Technical Evaluation Report, the tentative decision document, and
the NPDES [301(h)] permit should be reviewed to determine requirements for
sampling other biotic groups. If this review indicates the need to sample
phytoplankton, zooplankton, or fish communities, sampling locations should
be specified in the vicinity of the ZID boundary, other areas of expected
impact, and at one or more control sites. The number of control sampling
locations, as well as the need for and number of any sampling sites
specified in nearfield areas, should be determined based on the nature of
the potential impacts of the discharge. As in the example previously
presented, sampling depths should be specified for each of the above
sampling station groups. Determination of sampling depths must be made on
the basis of oceanographic conditions and behavioral characteristics of the
organisms; these depths may vary seasonally. The monitoring of habitats of
limited distribution may also be required in the nearfield area at a single
site or at several locations including control sites.
There would be additional station requirements for discharges into
stressed waters or in situations where other pollutant sources could
potentially affect biological communities in the vicinity of the applicant's
discharge. In such cases, it is important to define the magnitude of the
discharge interaction(s) and describe any biological response gradients
associated with the applicant's discharge and other pollutant sources in the
study area. Therefore, several additional stations may be required at
intermediate positions between the applicant's discharge and other
significant pollutant sources in the study area.
For cases where there is an improved discharge involving outfall
relocation, monitoring is required in the vicinity of the ZID boundary at
both the relocated discharge site and at the existing discharge site until
42
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discharge at the latter site ceases. Within-ZID stations, if necessary,
should be located close to the midpoint of the diffuser. ZID-boundary
stations should be oriented in the direction of the predominant current.
Single within-ZID and ZID-boundary stations should be sufficient for small
discharges, while two or more stations of each type may be needed for larger
discharges.
Selection of control stations is one of the more important aspects of
the design of the monitoring programs, since all assessments of impacts will
rely on comparisons made with data from these locations. Control stations
should be located outside the traceable waste field and not be affected by
the applicant's discharge. Similarly, the selected locations should not be
influenced by other discharges. Control stations should be located in water
of similar depth to that of the within-ZID, ZID-boundary, and gradient
stations. Sediment characteristics should be similar at all sampling
stations except where sediment alterations are due to an outfall effect.
Control and other monitoring stations should be located approximately the
same distance from shore. Since it is often necessary to locate control
sites a considerable distance [5-10 km (3-6 mi)] from the outfall to escape
all waste field influences, candidate control sites should be carefully
evaluated to ensure that oceanographic conditions are not atypical.
Example layouts of sampling locations for two alternative biological
monitoring programs are presented in Figure 1. Included in the example are
sampling stations that should be expected at a relatively large (x stations)
or small (o stations) discharge. Some of the important features common to
both layouts that should be noted are: sampling stations have been located
at the same depth and at approximately the same distance from shore;
near-ZID and nearfield gradient stations are positioned in the same
direction from the diffuser as the predominant current direction; and,
control stations are located a considerable distance from the diffuser and
in the opposite direction of predominant currents.
Although type(s) of stations are not specified in the figure, the
layouts are typical of expected benthic sampling designs. Sampling of other
biotic groups, if required, might also be conducted at these stations. The
number and location of stations are indicative of the most basic programs
that would be expected of relatively small and large discharges. In the
43
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NEARFIELD i« o x.
ZID
o Small discharge station
x Large discharge station
CONTROL
Figure 1. Representative sampling locations for two levels
of biological monitoring.
44
-------
case of a large discharge where a high potential for sediment accumulation
offshore has been identified, a number of stations would be required in the
deeper water offshore of the diffuser and at appropriate control stations.
These example layouts also do not reflect the existence of areas of special
concern, such as important fish habitats where additional sampling might be
required.
Sampling Frequency and Replication
The 301(h) regulations do not offer explicit guidance for either sample
frequency or replication. For those biological communities likely to be
affected by the discharge, sampling frequency will be dictated by
community-specific characteristics. For example, due to the rapid response
of phytoplankton to environmental perturbations and seasonal fluctuations in
community structure, the most effective sampling strategy might be intensive
sampling for relatively short periods of time. Similarly, the ability to
sample juvenile fish in nursery areas may be limited to certain seasons of
the year. These examples point to the fact that sampling strategies should
be considered in the sampling design. In the development of the strategies,
data should be reviewed carefully to consider life history characteristics
of target species.
Sample replication requirements are both site- and species-specific.
Decisions on the level of sample replication or sampling effort should
include careful consideration of the minimum detectable difference in
selected biological parameters. Field experiments should be planned
carefully in order to define minimum detectable difference levels, to
establish the number of replicate samples required, and to specify the
appropriate analytical approach.
Prior to designing a sampling program, the applicant should consider
two important criteria associated with the sensitivity of the study to
changes in biological parameters. These are the probability of rejecting
the null hypothesis when it is true (commonly called the probability, or
Type I error) and the probability of accepting a null hypothesis when it is
false (commonly called the probability, or Type II error). The complement
of e (1 - 3) is referred to as the power of a test and is especially
important since it defines the probability of correctly detecting
experimental effects (e.g., differences among sampling stations).
45
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For a specified variance associated with a biological parameter, the
probabilities of a, 3, and minimum detectable differences among sampling
areas can be expressed as a function of sample size. The allocation of
sampling resources (stations, replication, and frequency) can then be
determined with regard to available resources, practicality of the design,
and desired sensitivity of the subsequent analyses. Discussions and
examples of this approach are included in Cohen (1977), Winer (1971),
Scheffe (1959), Moore and Mclaughlin (1978), Gordon et al. (1980), and Sail a
et al. (1976).
Sample Collection and Processing
The following subsections provide a discussion of appropriate sample
collection, sample handling, and quality assurance/quality control methods
for the individual biotic groups which may be included in a 301(h)
monitoring program.
Benthos--
Most biological monitoring programs will emphasize the macrobenthos
since micro- or meiofaunal benthic samples are difficult and expensive to
process and also present interpretive difficulties due to extreme
small-scale heterogeneity and lack of understanding of community
relationships. Should impact upon micro- and meiofaunal benthos prove
significant for some outfalls such that monitoring of these infaunal
components is required, the investigators should consult Fenchel (1969),
Wieser (1960), Mclntyre and Murison (1973), and Hulings and Gray (1971) for
information on sampling methods and sample handling.
The methods and equipment for sampling macrobenthic infaunal
communities have been the subject of several publications [Holme and
Mclntyre (1971), Word (1976), Hedgpeth (1957), and Swartz (1978)]. The
ideal bottom grab for sampling all sediment grain sizes, from sand to silt,
has yet to be invented. Word (1976) compares the sampling efficiency of
seven grab samplers (Ponar, corer, Shipek, van Veen, orange peel,
Smith-Mclntyre, and a chain-rigged van Veen) in the silty-sand to
clayey-silt sediment off southern California. The results of Word's
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investigation suggest that a 0.1-m2 chain-rigged van Veen grab is the most
reliable sampling device for these sediment types. Swartz (1978) recommends
the use of a 0.1-m2 Smith-McIntyre grab due to its essentially constant
"bite" area; however, he recognizes that the depth of penetration of this
grab varies with sediment type.
Regional consistency in infaunal monitoring is important to the 301(h)
program objective. In the southern California area, for example, sampling
should consist of replicate sampling using a 0.1-m2 chain-rigged van Veen
grab. In more sandy areas, this grab or the Smith-Mclntyre grab recommended
by Swartz (1978) may prove acceptable; however, investigators studying sandy
infaunal communities should define the sampling efficiency of whichever grab
they chose to utilize.
In areas where visibility and oceanographic conditions permit,
diver-operated coring or dredging may be more desirable than grab sampling
from a surface vessel. The type and size of sampling device suitable for
each kind of substrate may vary from suction dredges (Brett 1964; Gale and
Thompson 1975) which cover large areas (for substrates with a low density of
organisms) to small coring tubes or small box corers (for substrates with a
fairly high density of infauna).
The number of replicate samples collected at each station should be
sufficient to ensure statistical reliability (see Sampling Frequency and
Replication above and Effect of Sample Size below). At each station, one or
more additional, separate sediment sample(s) should be collected for
analysis of total organic carbon content, grain size distribution, and
percentage of gravel, sand, silt, and clay. Other physical-chemical
parameters discussed in the water quality monitoring section of this report
should be monitored at or near each benthic station.
Sample Handling—The monitoring design should describe all procedures
used in the benthic sampling program. These descriptions should include the
following requirements:
1. Each replicate sample should be screened and preserved in
the field on the day of collection.
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2. Each replicate sample should be screened, fixed, sorted, and
processed separately.
3. Samples should be screened through a sieve having 1.0-mm
mesh. If smaller mesh size screens are also used (e.g., 0.5
mm), the fraction of the organisms retained by the 1.0-mm
sieve and smaller sieves should be processed separately.
4. Organisms should be fixed in a buffered 10-percent
formalin-seawater solution. (Borax is suggested as a
buffering agent.) The specimens should be transferred to a
70-percent ethanol solution after an initial fixation period
of 24 hours to 1 week. Vital staining techniques may be
used as an aid to sorting (see Holme and Mclntyre 1971,
Williams and Williams 1974).
5. Permanent labels should accompany each sample throughout all
phases of sample handling, processing, and storage. These
labels should include the date and time of collection, the
station and replicate identification number, the station
location including at least latitude and longitude, and the
sample collection depth. If available at the time of sample
collection, other label entries should include water
temperature, salinity, dissolved oxygen, and bottom depth.
Sample Processing—
1. The organisms should be sorted and identified to species,
or, if unidentifiable, sorted into discrete taxa.
2. The total number of individuals of each species (or lowest
identified taxon) in each replicate should be determined.
Counts should be expressed as the number of individuals of
each taxon in the samples and per m .
3. The wet weight of organisms in the six major taxonomic
groups (polychaetes, crustaceans, molluscs, coelenterates,
ectoprocts, and echinoderms) and the total biomass of each
replicate sample should be determined.
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Effect of Sample Size—The accuracy and precision with which benthic
community parameters are estimated depend on the parameter in question and
on the size of the sample. It is therefore appropriate to discuss the
effects of sample size on estimates of parameters most frequently used to
describe benthic community structure and function.
The total area sampled among the replicates at each station should be
large enough to estimate a given parameter within acceptable limits of
accuracy and precision. Within a study area, adequate sample size may be
determined by maximizing the number of species collected or by minimizing
the error of estimation of the mean for the parameter in question (Gonor and
Kemp 1978). Alternatively, the surface area sampled may be determined by a
review of sample sizes which in past studies have been shown to yield data
with acceptable limits of accuracy and precision. If the surface area
sampled per station is too small, the data will poorly estimate the
parameter in question because the ratio of the variance to the mean for a
given parameter will be unacceptably large (Gonor and Kemp 1978).
Consequently, within-habitat variability (which is a function of nonrandom
distribution of the fauna) will obscure differences in community structure
when stations are compared.
Holme and Mclntyre (1971) and Swartz (1978) recommend that an area of
0.5 nr (5.4 ft*) be sampled to assess species composition in coastal and
estuarine regions. This recommendation is supported by the results of
benthic studies in Puget Sound (Lie 1968). From an analysis of ten 0.1-m2
(1.1-ft^) replicates at one site, Lie concluded that a minimum of five
replicates is needed to accurately assess species composition, while a
minimum of three replicates is required to accurately estimate biomass and
numerical abundance.
Word (1976) presented an analysis of 10 replicate samples from southern
California (location not given). He observed that: 1) the cumulative
number of species does not appear to approach an asymptote with increasing
number of samples, and 2) a second sample will include newly acquired
species which constitute only 10 percent of the individuals in the first and
second samples. Word concludes that because numerical clustering strategies
are sensitive to species which contribute 90 - 95 percent of the total
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number of individuals, a single 0.1-m2 (1.1-ft2) sample is sufficient to
obtain "useful descriptive information" (Word 1976).
Although Word et al. (1980) has shown that a single 0.1-m2 (1.1-ft2)
sample is appropriate to describe the Infaunal Trophic Index (ITI is a
single number characterizing the trophic organization of soft bottom benthic
communities) in southern California, the single sample limits the degree of
community characterization. With only a single sample, there is no direct
estimate of within-group variance for statistical analyses. Because
individuals are distributed logarithmically among the species of a community
(Preston 1948, Sanders 1968), the species collected in the second and
successive replicates most often will be numerically "rare." Note that
"rare" is not synonymous with "unimportant." Predators, for example, are
usually "rare" because they are one trophic level removed from their prey;
yet, predators are usually a major factor influencing the diversity,
structure, and function of benthic communities (e.g., Connell 1961, Paine
1966, Bilyard 1974). Hence, it should be acknowledged that one 0.1-m2
(1.1-ft2) sample is generally not adequate to characterize benthic community
structure and function. Many uniformly distributed "rare" species which are
important in maintaining community structure and function will not be
captured in a single sample. In general, then, five replicate samples per
station are recommended for determining benthic community structure and
function, unless evaluation of site-specific data indicates that sufficient
sensitivity could be obtained with fewer samples or that a greater number is
required due to extreme spatial heterogeneity.
The previous discussion concerns the number of replicates (or area
sampled) generally required to adequately characterize infaunal communities.
The other major aspect of sample size concerns the statistical sensitivity
or power associated with the number of replicate samples. A discussion of
the statistical aspects is included in a previous section (Sampling
Frequency and Replication).
Quality Assurance and Control Procedures—To assure proper handling and
processing of benthic samples, the following procedures are recommended:
1. At least 5 percent of all samples should be resorted by
individuals different from those who conducted the original
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sorting. Records on the results of this second sorting
should be maintained and presented in an appendix to the
monitoring report. This should be a double-blind test.
2. Complete sorting, processing, and/or laboratory records for
each replicate sample should be included in a separate
appendix volume to the annual report. These record sheets
should present as a minimum the data specified in Item 5 of
Sample Handling (above). The names and detailed statements
of the qualifications of all persons performing and
confirming taxonomic identifications of organisms should be
included in the appendix volume.
3. A voucher collection consisting of specimens representative
of each species (or lowest taxonomic unit of identification)
collected during this monioring program should be developed
and maintained by the applicant. This collection should be
archived for a period of not less than 2 years after the
expiration date of the 301(h) modification and should be
housed at a facility where adequate curatorship can be
assured.
4. Taxonomic references used for the identification of
organisms should be cited in the appendix to the report.
Bioaccumulation Studies—
The amended 301(h) regulations require periodic determinations (except
for small applicants meeting certain depth and solids deposition criteria)
of the accumulation of toxic pollutants and pesticides in organisms and
examination of other adverse effects of the discharge such as disease,
growth abnormalities, physiological stress, or death. At discharges where
bioaccumulation of toxic substances is known or likely to be a problem,
tissue samples from resident macroinvertebrates and fish species should be
examined for abnormal body burdens of toxic substances. The identified
toxic pollutants to be monitored are the 129 priority pollutants plus six
pesticides listed in Table 3.
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The primary method to be considered for determining levels of toxic
substance bioaccumulation in the vicinity of the outfall should be through
the use of caged specimens of bivalve molluscs. Recommended methods are
provided by EPA (1982). In situ biomonitoring has been used to monitor
levels of toxic substances in the water column (Young et al., 1976).
Generally, mussels (Mytilus californianus or M. edulis) or oysters
(Crassostrea spp.) should be utilized as the test organisms. These species
are widely distributed and easily collected in large numbers in most coastal
areas of the U.S. Also, there is a considerable amount of literature
concerning the rate of uptake of specific substances, experimental survival,
and the selective uptake of the various groups of toxic substances by
different tissues (see, for example: de Lappe et al. 1972; Young and Heesen
1974; Clark and Finley 1973; Alexander and Young 1976; and Eganhouse and
Young 1976). Other filter-feeding molluscs have been used in a similar
manner to monitor toxic substances in the marine environment (Goldberg et
al., 1978), and these organisms could be substituted if conditions are not
appropriate for survival of mussels or oysters.
Although minimum numbers of replicate samples and specimens are
specified in EPA (1982), the investigation of site-specific environmental
characteristics, seasonal variability in background pollution concentration
levels, and, most importantly, variability in the uptake of different toxic
pollutants should be examined during the initial stages of the
bioaccumulation monitoring program. The objective of these preliminary
investigations should be to determine the number of replicate samples and
number of organisms included in composite samples which will result in the
optimal sampling program, i.e., a program that will provide the basic
assessment information at a minimum cost for sample collection and analysis.
The monitoring of toxic pollutant concentrations in tissue samples from
resident macroinvertebrates and fish species should be conducted in cases
where the bioaccumulation of toxic pollutants has been documented or the
potential for accumulation in the food chain is considered to be high.
Emphasis should be placed on the selection of commercially or recreationally
important species for which information is available on the uptake and
effects of elevated toxic pollutant concentrations. Composite samples
consisting of at least six specimens should be collected at the specified
stations and sampling periods for tissue analysis. Since the objectives of
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this phase of the program are to investigate accumulation of toxic
pollutants in the food chain and to assess the suitability of commercially
or recreationally important species for human consumption, emphasis should
be placed on the determination of toxic pollutant concentrations in muscle
tissue.
At the time of sample collection, the length, weight, and mortality (in
the case of the caged bivalves) should be recorded. The physiological
condition of all organsims (e.g., the presence of external lesions and
discoloration) should also be noted. The observed concentrations of
identified toxic pollutants in tissue samples should be reported in both
tabular and graphic form. Statistical comparisons of the observed
concentrations of toxic substances in tissue samples should be made to
determine the existence of significant differences among stations or
replicates. Degree of fouling of cages and presence of potential predators
(e.g., crabs) within cages should be noted.
Fishes--
Marine fish communities are complex and dynamic in nature. The
structure of these communities changes seasonally as a result of spawning,
migrations, and recruitment of juvenile fish to adult populations. In the
short term, feeding activities (including diel movements) will influence
observed community composition. Selective characteristics of various types
of fishing gear tend to confound this problem (Hamley 1975) since, for
example, different sized (and aged) individuals from the same species will
be selected differently by different gear types, while individuals of the
same size from different species will also exhibit differences in
catchability. Catchability also varies on a diel basis as a result of
changes in avoidance capabilities under different light conditions, effects
of tidal currents on activity patterns, and other factors.
Extreme spatial heterogeneity is a characteristic aspect of the
distribution of many species of demersal and (especially) pelagic species;
sampling plans which fail to take spatial heterogeneity into consideration
can result in biased conclusions.
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Spatial and temporal variation of this kind places significant demands
on the design of sampling surveys. For example, Richkus (1980) reports
that, in a study conducted by Texas Instruments, Inc., in Chesapeake Bay, it
was determined that it would be necessary to collect 252 samples to produce
an 80 percent chance of detecting a 50 percent difference in density of
white perch (Morone americana) among three locations. Thus, the objectives
of a monitoring program for fishes, together with the structure of the
target community, will exert a major influence on many aspects of the design
of that program.
Several different types of gear have been used for sampling fish
communities. The selection of gear which is appropriate to address specific
survey objectives will depend upon the substrate type, the communities to be
sampled, tidal and other current conditions, depth, proximity to the shore,
and survey vessel capabilities (Table 7). A discussion of the applicability
of various fish sampling techniques is included in Richkus (1980). Von
Brandt (1972) provides a general review of fish catching methods. Uzmann et
al. (1977) present a comparison of three survey techniques.
When demersal fish populations are to be sampled in areas of sand or
mud bottom, use of an otter trawl is appropriate. The Marinovich 7.62-m
(25-ft) headrope otter trawl, described as the Coastal Water Project
Marinovich net in Table 3 of Mearns and Stubbs (1974), is recommended for
this purpose. It is commonly used for environmental survey work and is
easily handled from a small boat (Mearns and Stubbs 1974). The net is towed
from a single warp. Gear specifications and sampling procedures are
critically important. A discussion is provided by Mearns and Allen (1978),
but several additional points are important:
• Steel (or stainless steel) towing cable of 6.35 mm (0.25 in)
minimum diameter is recommended.
• A power winch is required; gear must be recovered while the
vessel is moving forward.
• Effort must be reported in terms of distance or area
covered. Fixed buoys or navigational aids (e.g., the Mini
Ranger) will be useful in this context. Haul distances of
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TABLE 7. SELECTION OF APPROPRIATE FISH SAMPLING GEAR
Demersal
Pelagic
Nearshore
Primary Approach
Otter Trawl
(Gillnet,
Trammel Net,
or Trap)3
Commercial
Monitoring,
Gillnet
Beach Seine
Habitat
Secondary Approach
Diving
Submersible
Hook and Line
Acoustic Transect
Pelagic Trawl
Lampara/Purse Seine
a If not trawlable.
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700 - 1,000 m (2,297 to 3,281 ft) are recommended.
Information on vessel speed and haul duration should be
recorded but cannot substitute for distance estimates.
• A towing speed (relative to the bottom) of 1.3 m/sec (2.5
knots) is recommended.
• The gear should be towed into the current and along an
isobath. Current conditions should be recorded.
A fundamental problem of demersal trawl sampling concerns the manner
with which this gear samples pelagic forms. These species (or life history
stages) often exhibit schooling behavior. Incidental encounter of pelagic
forms occurs during setting and recovery of gear; this is inconsistent, but
individual species behavior is not always understood well enough to permit
objective exclusion of these data. Species that are unquestionably pelagic
in habit (such as the northern anchovy Engraulis mordax) should be excluded
from demersal trawl catch data. This problem can also be addressed by
separately analyzing data for one segment of the fish community which is
known to be demersal (such as the flatfishes). An awareness of the
selective characteristics of trawl gear and the behavior of the gear itself
is also important (Wathne 1977, Harden Jones et al. 1977).
Gill nets and trammel nets are often utilized in areas where bottom
conditions preclude trawling or where improved spatial resolution is
required (such as within the ZID and at comparable stations). Variable
mesh, set gill nets are recommended (Ricker 1971), although trammel nets
(Becker et al., 1975) may be appropriate in some situations. Traps also
provide high spatial resolution (Becker et al., 1975) but are highly
selective.
In situations where nearshore fishes need to be sampled (such as when
an outfall is located in an area of juvenile salmonid migration), beach
seine sampling should be conducted (Allen et al., 1959).
Pelagic forms are often ignored during survey sampling. Pelagic
species are, however, important in many of the areas for which 301(h)
applications have been received. In a situation where the species of
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concern are subject to consistent commercial or recreational exploitation in
the vicinity of the outfall, fisheries surveys (see below) can be utilized
to collect appropriate data. When this is not possible, the use of acoustic
(and sonar) transects and pelagic sampling nets (such as pelagic trawls,
lamparas, and purse seines) is recommended (Saville 1977; Becker et al.,
1975; Lemberg 1978; Fiedler 1978; Richkus 1980).
Specific data can be collected from the immediate vicinity of the
outfall structures by means of diving and submersible surveys. Details are
provided by Allen et al. (1976), Allen (1975), and Becker et al. (1975).
Line transect techniques often provide an appropriate method for quantifying
diving observations; quantitative procedures are discussed by Seber (1973).
Some field methods are presented by Fager et al. (1966) and Walton and
Bartoo (1976).
Hook-and-line surveys are especially useful for sampling in precise
locations and for obtaining larger individuals which may avoid other
sampling gear. This type of gear can be used in areas which may not be
accessible to other sampling devices and is frequently employed to provide
specimens for bioaccumulation analysis. Allen et al. (1975) describe
appropriate techniques. It should be noted that hook-and-line techniques
are especially selective in nature; hook size, bait type, and other gear
specifications should be selected with this consideration in mind.
Data collection requirements will depend on specific survey objectives.
At a minimum, all fish catches should be identified to species, and counted
and weighed by species. Taxonomic procedures and authorities should be
clearly defined and a procedure for seeking expert advice should be
included, if specimens cannot be identified by employees or consultants.
For individual species, length-frequency and length-weight analyses may
be required to allow consideration of population differences between outfall
and reference sites. Standard length is the recommended measurement; if a
different length measurement is recorded, this should be stated, and a
regression relationship between this measurement and standard length should
be provided together with the data utilized in the analysis. Appropriate
subsampling procedures must be defined for the collection of these data.
When individual observations are recorded, life history stage should also be
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reported. Unless subsampling is to be conducted, all individuals in the
catches should be examined for external disease symptoms or abnormalities; a
standardized technique for identifying and defining abnormalities should be
developed and included in the data reports. Each individual observation
should be recorded; the use of computer data coding forms is recommended for
this purpose. All raw data observations, identified by sample number,
station location, date and time of collection, and individuals responsible,
should be included as an appendix to reports.
Commercial and Recreational Fisheries—
If commercial or recreational fisheries activities are conducted in the
vicinity of an outfall, these activities must be monitored. Commercial and
recreational fisheries catch data are generally reported as summary
statistics for statistical blocks defined by the state agencies concerned.
In most cases, the area covered by these statistical blocks is too large to
allow resolution of fishery catch conditions close to a sewage outfall.
Coordination with state agencies may provide an effective and
inexpensive mechanism for collecting data which can be used to assess the
condition and productivity of specific fisheries. A voluntary logbook
program for fishermen could be designed which would allow those fishing in
areas of concern to record data on catch and effort close to the outfall and
at remote locations. In some cases, vessels observed to be fishing close to
an outfall could be identified and the operators Interviewed when the
vessels dock. Similarly, individuals observed sportfishing could be
identified and interviewed.
An alternative approach to monitoring recreational fishing activities
would involve sampling at the outfall and reference area with appropriate
sportfishing gear. This would allow direct comparison of species caught,
catch rates, disease prevalence, bioaccumulation, and other relevant
aspects.
Market or consumer acceptability of fish caught in the vicinity of a
POTW outfall should also be addressed in all commercial and recreational
fisheries surveys; a simple interview procedure would be appropriate.
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Recreational harvesting of intertidal shellfish resources occurs in the
vicinity of some POTW outfalls. These shellfish should be sampled with
respect to both public health considerations and possible population
responses to discharge effects. While quadrat sampling (Seber 1973) is a
theoretically attractive approach to this problem, the time and effort
involved would be prohibitive in most cases. Monitoring of recreational
catch and effort (in association with the appropriate government agency) is
an appropriate way to examine relative population abundances.
When interviews, voluntary logbooks, and field observations of fishing
activity are utilized for data collection, a complete log of all relevant
information should be maintained by the POTW. Interviews should be
conducted by means of questionnaire. The log should contain all interview
and logbook returns and detailed records of field observations; details of
public health analysis of shellfish will also be included. All entries
should be identified by time, date, and the individuals who collected the
information. A clear and comprehensive copy of this log should be included
as an appendix to monitoring reports.
Phytoplankton--
Since phytoplankton are transient, a monitoring program to sample
phytoplankton should be designed somewhat differently from monitoring
programs for certain other biotic groups. Phytoplankton are carried about
by movements of the water, and consequently maximum sewage effluent impacts
on phytoplankton may occur well away from an outfall. Stations should be
located at sufficient distance from an outfall to accommodate a lag time in
the response of phytoplankton to sewage effluent inputs. Due to their short
turnover times (on the order of hours to days), phytoplankton communities
may respond to perturbations much more rapidly than other biotic groups;
therefore, samples must be collected more frequently. Bimonthly samples are
probably the least frequent which could be expected to give meaningful
results, although monthly or even biweekly samples would be preferable.
In situations where phytoplankton communities display pronounced
seasonal variations in standing stock or production, it may be appropriate
to use a temporally stratified sampling approach. For example, if
phytoplankton growth is highest during the spring, sampling may be conducted
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on a frequent basis (e.g., weekly). Similarly, during periods of
consistently low phytoplankton growth, a reduced sampling frequency may be
used.
Adequate assessment of phytoplankton community response to sewage
discharges will generally involve more sampling stations than would be
required for the benthos. Thus, it is important to initially assess whether
or not a phytoplankton sampling program is justified based on a
consideration of discharge size, sensitivity of receiving water, and
evidence of previous impacts on phytoplankton. Selection of phytoplankton
sampling stations should always involve a thorough consideration of water
circulation patterns to ensure that putative waste field stations are
actually being exposed to the diluted effluent and that control stations are
not subject to influence of the waste field. If evaluation of circulation
patterns indicates that the sewage waste field may be transported to areas
of limited flushing (e.g., embayments or eddies), special emphasis should be
placed upon locating sampling stations in these areas. In all cases,
phytoplankton sampling stations should be located in areas of maximum
predicted effects, considering such factors as response lag time, effluent
dilution, and circulation patterns.
The most likely direct effect of sewage effluent on phytoplankton
communities is enhancement or inhibition of primary production. Enhancement
may occur in areas where the phytoplankton are naturally nutrient limited,
since sewage effluent represents a significant source of nutrients.
Inhibition may occur if there are sufficient concentrations of toxic or
inhibitory substances in the effluent.
In areas where phytoplankton production is enhanced (or inhibited), the
standing stock of phytoplankton may be expected to be higher (or lower) than
in reference areas. Measurement of the concentration of chlorophyll a in
the water is an indirect method of analyzing the standing stock of
phytoplankton. It is recommended that the two-dimensional spatial
distribution of chlorophyll ^ concentrations about the outfall be analyzed.
Samples should be collected at several distances from the outfall in the
direction of current flow. Samples should also be collected at a variety of
depths throughout the euphotic zone (from the surface to the 1-percent light
level, as estimated from any light transmittance data collected as part of
the Water Quality Monitoring Program).
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The vertical distribution of chlorophyll a^ concentrations at each
station may be examined through the collection of water samples with water
bottles at various depths (followed by fluorometric or spectrophotometric
determination of chlorophyll a) or, if available, a pump system may be used
with a flow-through fluorometer (Lorenzen 1966) for a continuous profile of
chlorophyll ^concentrations vs. depth. The advantages and disadvantages of
various water bottles are discussed by Venrick (1978a), while the use of
pumps is discussed by Beers (1978). It is advisable that a pump be used
only for the determination of chlorophyll a^ and that water bottles be used
for the collection of phytoplankton for productivity measurements and
taxonomic analyses, since the inevitable agitation associated with pumping
may damage some cells.
If it is determined that the vertical distribution of phytoplankton
biomass (as mg chlorophyll a/nr*) is reasonably uniform throughout the
euphotic zone, water samples for simulated in situ primary productivity
measurements (Ahlstrom 1969) may be collected with water bottles at depths
corresponding to fixed percentages of incident solar radiation. If,
however, there is significant vertical stratification of the phytoplankton
community, sampling depths should be adjusted so that samples are also
collected within subsurface chlorophyll maxima or minima. Phytoplankton
primary productivity should be measured by the 14c light-dark bottle
technique as described by UNESCO (1973), and measurements at each
station-depth should be replicated to facilitate statistical analysis.
If the monitoring program described above reveals perturbations of
chlorophyll £ concentrations and/or primary productivity within or beyond
the ZID, taxonomic analyses should be conducted since phytoplankton species
vary in their responses to alterations in their nutrient source (Eppley et
a!., 1969) or in their responses to certain inhibitory substances (Thomas
and Seibert 1977). Subsamples should be drawn from the water collected in
the water sampling bottles and preserved for later microscopic analysis
onshore. It is important that samples for taxonomic analysis be collected
at various depths throughout the euphotic zone since different species may
have different depth distributions. It is also important that sampling be
conducted at similar times during the day (i.e., mid-morning or
mid-afternoon) since some phytoplankton are known to migrate vertically
(Stofan and Grant 1978).
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Choice of a fixative depends somewhat on the dominant types of
phytoplankton known to inhabit a given area. Buffered formaldehyde and
Lugol's solution are two common fixatives. The advantages and disadvantages
of each are discussed by Throndsen (1978a).
Taxonomic analysis almost always involves some form of subsampling,
which is consequently a potential source of bias or variability. The
statistical implications of subsampling are discussed by Venrick (1978b).
Preserved phytoplankton samples normally must be concentrated for
quantitative microscopic analysis. Although other methods are available
(Sukhanova 1978, Throndsen 1978b), the routine method is the Utermohl
technique, which utilizes sedimentation cylinders and an inverted microscope
(Hasle 1978).
Taxonomic analysis should include identification of the dominant
phytoplankton taxa and counts of individual species whenever possible.
Numerous taxonomic references are available [see Chapter 6.4 of UNESCO
(1978)]. The accuracy and consistency of phytoplankton identifications are
of the utmost importance for characterization of the BIP, both in the
reference area and at stations in the vicinity of the discharge. Counts of
individual species should be standardized to numbers per unit volume of the
original water-bottle sample, calculated with consideration for whatever
subsampling technique was utilized.
If replicate taxonomic samples are available for each station-depth,
the estimates of abundance of individual species may be analyzed
statistically for differences among depths, among stations, or among times.
Particular attention should be given to differences in community composition
between stations in the vicinity of the outfall and stations in a reference
area. Species diversity, richness (number of species), evenness, or
numerous other parameters (Pielou 1970) may be utilized for description and
comparison of the phytoplankton communities.
If available information indicates a potential for enhancement of
individual phytoplankton groups (especially dinoflagellates), the monitoring
program should include an assessment of the magnitude, duration, and point
of initiation for phytoplankton blooms. Special emphasis should be placed
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upon species causing accumulation of toxins in other organisms, or blooms
which may result in fish kills.
The goal of the phytoplankton monitoring program should be to
demonstrate whether or not the discharge of sewage effluent from the outfall
in question interferes with the protection and propagation or the natural
range of variation of phytoplankton in areas beyond the ZID.
Characteristics of phytoplankton which may be examined include the community
biomass (as estimated through the measurement of chlorophyll £
concentrations), community primary productivity (as estimated through
simulated in situ incubations using the light-dark bottle technique), and
the various community composition parameters. The responses of biological
communities to pollutant stress appear to involve a continuum, as indicated
by the gradients in the biological variables of the benthos near sources of
organic pollutants (Pearson and Rosenberg 1978). Therefore, alteration to
the phytoplankton communities should be analyzed in relation to potential or
determined (by other community studies) impacts on other biological
communities which make up a balanced indigenous population. This analysis
should include, although it should not be limited to, food web impacts, the
occurrence of toxic or nuisance phytoplankton, eutrophication or blooms, and
potential second impacts on zooplankton or fish communities.
Zooplankton—
Zooplankton, like phytoplankton, are transient, and, consequently, a
monitoring program designed to sample zooplankton should be designed
somewhat differently from monitoring programs for certain other biotic
groups that tend to be permanent residents of an area. Zooplankton are
carried about by movements of the water; therefore, the maximum sewage
effluent impacts on zooplankton may occur at some distance away from the
outfall. Unlike phytoplankton, however, zooplankton life spans are
typically on the order of a few months, so their capacity for responding to
perturbations is much less than that of phytoplankton. Bimonthly samples
are usually adequate for analysis of changes in zooplankton communities.
Zooplankton possess varying degrees of swimming ability and therefore have
the potential for aggregating in patches or in narrow depth strata, which
introduces an additional complication in quantitative sampling. In
addition, the ability to swim means that many zooplankton can avoid certain
types of sampling gear.
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As is the case for phytoplankton, the design of zooplankton sampling
programs should consider natural temporal fluctuations in abundance and
species composition. Zooplankton assemblages display a high degree of
spatial heterogeneity in addition to pronounced diel vertical migrations by
many groups. These factors, combined with a longer response time to effects
of sewage effluents, would result in the necessity of conducting relatively
extensive programs (i.e., in number of sampling stations, frequency of
sampling, and replications) to adequately assess responses of zooplankton to
pollutant inputs. Thus, studies of zooplankton assemblages should be
conducted only when there is evidence of previous impact in zooplankton,
when phytoplankton communities display significant effects, or when large
discharges are located in areas where there is a high potential impact on
zooplankton (e.g., in estuarine environments with important macroplanktom'c
larvae of commercial and recreational species).
For zooplankton, there is no easily measured functional response to
POTW effluent discharges similar to primary productivity for phytoplankton.
Toxic effects of effluent on zooplankton are possible if there are
sufficient concentrations of toxic substances in the effluent. Alteration
of zooplankton community composition is a distinct possibility in areas
where the phytoplankton community composition has been affected, since many
zooplankton graze on phytoplankton. Given the smaller proportion of their
life spans spent within the sphere of influence of the outfall, zooplankton
are less likely to experience direct, observable changes in community
composition than are phytoplankton.
The zooplankton encompasses a wide range of organisms, from microscopic
protozoans to large planktonic crustaceans and larval fish. In general, the
smaller organisms have shorter life spans, so effluent impacts are more
likely among the smaller organisms. Sampling methods vary depending upon
the size of the organisms.
Microzooplankton (those which pass through the mesh of a 202-um net)
can be collected either with water bottles similar to those used for
phytoplankton collection [although a volume of at least 10 liters is
recommended (Jacobs and Grant 1978)] or with pumping systems (cf., Beers et
al., 1967). If water bottles are used, samples should be collected from a
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variety of depths throughout the water column, and the captured organisms
may be concentrated with a fine-mesh (e.g., 63-um) screen. Replicate
samples should be collected from each station-depth. Pumping systems may
incorporate a single filter or a mesh of different size filters for
collection of the organisms. Flow rate should be in the range of 150 - 200
1/min (Jacobs and Grant 1978). Pumping systems have the advantage of being
able to take samples integrated over depth, or of collecting samples while
the ship is underway, but they may damage soft-bodied organisms, and they
are more expensive and complicated than water bottles.
For collection of small mesozooplankton (those retained on a 202-um
mesh), nets are generally used. For an excellent discussion of net design
and function, see UNESCO (1968). Nets with small mouth diameters (20 - 40
cm) may introduce error by underestimating abundance and diversity (McGowan
and Fraundorf 1966, Wiebe and Holland 1968). A minimum mouth diameter of 60
cm is generally recommended (Jacobs and Grant 1978). With mesh sizes of
less than 202 urn, clogging and loss of filtration efficiency are often a
problem, so a 202-um mesh is the smallest which should be used (UNESCO
1968). Additional tows may have to be made with larger nets (1.0-m mouth
diameter and 505-um mesh) in order to collect representative samples of
larger zooplankton and larval fish. All tows should be replicated; the
number of replicates necessary for the desired precision of estimates should
be determined during a preliminary or pilot sampling program (cf., Cochran
1963; Green 1979). Paired bongo nets (McGowan and Brown 1966) are often
used because they provide two replicate samples from the same environment.
Alternatively, they can be rigged with two di f ferent mesh nets for
collection of different size fractions. Any net used should have a flow
meter attached to the mouth for calculation of volume filtered. Prior to
selection of a sampling technique, the spatial and temporal distributional
characteristics of the target zooplankton assemblage should be considered.
For example, lobster (Homarus americanus) larvae occur near the water
surface and are appropriately sampled by a neuston net.
Oblique tows are highly recommended, with sampling extending from just
above the bottom to the surface. Avoidance by larger zooplankton is
significant at slow tow speeds, so a ship's speed of 1.5 - 2 knots should be
maintained (Jacobs and Grant 1978). The animals collected should be washed
(from the outside of the net) into the cod end, transferred to a labeled
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sample jar, and preserved with buffered formalin in the proportion of nine
parts sample to one part formalin (a saturated solution containing 38- to
40-percent formaldehyde). Further discussion of shipboard handling of
samples can be found in Jacobs and Grant (1978).
For quantitative taxonomic analysis of the zooplankton samples,
subsampling will normally be required. Two methods which are commonly used
are: subsampling by Stempel pipette, and splitting with either a Folsom
splitter (McEwen et al., 1954) or the newer Burrell et al. (1974) device.
In either case, large and/or rare organisms should first be counted and
removed. The use of the Stempel pipette is discussed by Jacobs and Grant
(1978), who point out that due to the small aliquot size, this method should
only be used when rapid "ballpark" numbers are needed. The use of plankton
sample splitters is also discussed by Jacobs and Grant (1978), and they
indicate that this is the best method for quantitative analysis of
zooplankton. The counts of individual taxa should be transformed to numbers
per sample (considering the subsample size) and then standarized to numbers
per unit volume of water filtered (calculated from the flowmeter reading).
Taxonomic analysis of the samples should include identification of the
dominant zooplankton taxa and counts of individual species whenever
possible. Particular attention should be given to the meroplanktonic larvae
of commercially and ecologically important species (e.g., fish, shrimp,
lobsters, etc.). The accuracy and consistency of zooplankton
identifications are of the utmost importance for characterization of the
range of variation of the zooplankton communities, both in the reference
area and at the various outfall stations.
If replicate taxonomic samples are available for each station, the
estimates of abundance of individual species may be analyzed statistically
for differences among stations or among times. Particular attention should
be given to differences in community composition between stations in the
vicinity of the outfall and stations in a reference area. Species
diversity, richness (number of species), evenness, or numerous other
parameters (Pielou 1970) may be utilized for description and comparison of
the zooplankton communities.
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The goal of the zooplankton monitoring program should be to demonstrate
whether or not the discharge of sewage effluent from the outfall in question
interferes with the protection and propagation or the natural range of
variation of the zooplankton communities in areas beyond the ZID.
Characteristics of the zooplankton communities should include, but not
necessarily be limited to, species composition, abundance, dominance, and
diversity. The responses of biological communities to pollutant stress
appear to involve a continuum, as indicated by the gradients in biological
variables of the benthos near sources of organic pollutants (Pearson and
Rosenberg 1978). Therefore, alteration in the zooplankton communities
should be analyzed in relation to potential or determined (by other
community studies) impacts on other biiological communities which would make
up a balanced indigenous population. This analysis should include, although
it should not be limited to, the structure and function of both larval and
adult zooplankton communities, as well as consideration of food web impacts.
Kelp Communities--
Kelp beds are distinctive habitats of limited distribution whose
protection is of special concern because of their ecological significance
and their economic value to man. The kelp plants themselves are largely
responsible for the spatial structure of this community, as they provide
food, substrate, and shelter for a variety of organisms (Tegner 1980). In
some areas, the kelp itself is harvested, and in many areas the kelp beds
are the location of valuable fisheries for abalone, lobster, fishes, and sea
urchins (Tegner 1980). Kelp beds may be particularly sensitive to outfall
discharges, and adverse effects of sewage effluent on kelp have been
suggested by Carlisle (1968), Mearns et al. (1977), and others. If kelp bed
communities are potentially affected by a sewage effluent discharge, a
monitoring program should be conducted to evaluate the health and extent of
these communities.
Kelp bed communities typically include a great variety of plant and
animal species, but since the continued existence of the community is
largely dependent on the presence of the kelp plants themselves, the
monitoring program should focus on the health and spatial distribution of
these plants, rather than attempt a detailed analysis of the entire
community. The location, condition, and size of kelp beds along the
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southern California coast have been monitored for a number of years through
the use of aerial surveys (Wilson et al., 1980). This is a particularly
useful technique because changes in the area! extent of the beds can be
monitored over time, and areas of potential sewage impact can be identified.
Aerial surveys of kelp beds utilize infrared photographs taken from an
aircraft flying at an altitude of 1.5 - 2.6 km (0.9 - 1.6 mi). Photographs
taken around midday will minimize reflected glare, and a polarizing filter
may be used, if necessary. Overlapping adjacent photographs (10 - 20
percent) assure full coverage and minimize barrel distortion at the film
edges. Slides of the kelp beds can be projected and drawn onto charts of
the coast, and the surface area of the kelp canopies can be calculated from
these charts using a polar planimeter or a measured grid network (Wilson et
al., 1980).
If a decline in nearby kelp beds occurs, the discharge of sewage
effluent may or may not be the cause or a cause. Sewage effluent may
adversely impact kelp in a variety of ways. Certain constituents of the
effluent may be toxic to kelp sporophytes and/or gametophytes, and the
plants may die in areas where destructive concentrations of these
constituents appear. Turbidity within the discharge field of the outfall
might be increased, reducing light intensities or altering the spectral
distribution of the light such that kelp growth is adversely affected.
Concentrations of kelp enemies such as grazers, pathogens, or parasites
might in some way be enhanced by the effluent, and the increase in their
numbers might bring about a decline in the kelp. Finally, siltation effects
from the settling of suspended matter discharged by the outfall might
interfere with the recruitment of young kelp plants (North 1964).
There have been suggestions (Wilson et al., 1980) that sewage effluent
discharge may have decreased the maximum depth of kelp growth in nearby kelp
beds. If the light transmittance data collected as part of the water
quality monitoring program is also collected on a regular basis along the
seaward edge of nearby kelp beds, it should be possible, given the known
photosynthetic requirements of kelp (Clendenning 1964; Rosenthal et al.,
1974), to determine whether the sewage effluent discharge may be adversely
impacting the kelp beds. Comparison of light transmittance measurements at
specific depths along the kelp bed in question with those at similar depths
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near other kelp beds removed from anthropogenic sources of sedimentary
material should indicate whether the kelp bed in question could extend into
deeper water in the absence of the effluent discharge.
The use of sediment traps to quantify the amount of sedimentation
occurring along the margins of potentially impacted kelp beds should also be
considered. Comparison of the rates of sedimentation there with those along
kelp beds removed from anthropogenic sources of sediments, and consideration
of the effects of different amounts of sediment on the survival and growth
of kelp germling stages (Devinny and Volse 1978), may permit an evaluation
of whether or not the sewage effluent discharge may be inhibiting expansion
of the kelp bed in question.
While the toxicity of certain effluent constituents on kelp has been
studied (North 1964), it is probably unreasonable to expect that detailed
studies of toxic effects would be conducted as part of a kelp monitoring
program, given the large number of potentially toxic or inhibitory
substances in most municipal effluents and the possibility for complicated
synergistic effects.
Increased abundances of certain animals which graze on kelp [notably
sea urchins (see Lawrence 1975)], have been implicated in the decline of
certain kelp beds. While there have been suggestions that the abundances of
these grazers may be enhanced by the discharge of sewage effluent (Clark
1969), it is difficult to establish cause-and-effect relationships. The
true cause of the increased abundances of these organisms may only be
revealed through detailed investigations of interspecific interactions and
predator-prey relationships (Tegner 1980), which may be beyond the scope of
individual 301(h) monitoring programs.
One promising method which may be used to infer causes of observed
changes in kelp canopy size is regression analysis, utilizing such factors
as suspended solids, mass emission rates, water temperature, water
transparency, etc. Wilson et al. (1980) used this method to examine
potential causes of the initial decline and subsequent recovery of kelp beds
along the Palos Verdes Peninsula, in close proximity to a very large sewage
outfall. Nevertheless, identifying changes which have occurred in kelp
canopy size is somewhat easier than deciding what factor(s) may be
responsible.
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Coral Communities--
In some areas, it will be important to assess any impact the
applicant's discharge may have on coral communities. Generally, an
assessment of changes on living coral coverage which may include a study of
reef fishes, will be sufficient rather than a complete study of the reef
community.
A line transect method of sampling should be used for studies of both
the living coral coverage and reef fishes. All stations should be
comparable as far as the distance from sand areas, rubble, and base-rock
relief. At each station, a 50-m (164.1-ft) length of electrical wire should
be permanently attached to the reef, parallel to the shoreline. Care should
be taken to ensure that the line is located within a reef area of sufficient
size so as to eliminate any patchiness in the data due to sand area or reef
edge effects.
Photographs of at least 0.67 m2 (7.24 ft2) of the bottom should be
taken on the shoreward side of the line at 5-m (16.4-ft) intervals. An
underwater camera mounted on a rigid framing device should be used. Each
photograph should contain a small slate indicating the station, date, and
position of the photograph along the line. Care shall be taken in order to
be certain of photographing the same quadrat each quarter. It is suggested
that stakes be driven or cemented to the reef indicating at least two
corners of each frame.
The photographs should be developed as slides. These slides should be
projected onto a grid having the dimensions of the original quadrat, and the
percent coverage of coral and encrusting algae by species, and of the
noncoral substrate, should be estimated.
Other transecting methods may be used to sample living coral coverage
if they are shown to be statistically valid sampling techniques; however,
due to the relative ease of sampling and data reduction, the photographic
method described above is recommended.
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The investigation of reef fishes should consist of SCUBA transects
along the transect lines used for the reef benthos survey. Beginning at
least 1 day after the lines are permanently attached to the reef, a diver
should swim each line in order to identify and count the fishes located
within 3 m (9.8 ft) either side of the line. The diver should take care to
enter the water away from the transect area in order to avoid disturbing the
fishes. One transect at each station should be completed during each of
three consecutive days for each quarter.
Consistency of technique is important; therefore, the applicant should
make every effort to ensure that transects are conducted in a similar manner
by the same diver-biologist if possible. Some investigators are known to
look for cryptic species more than others, or to notice larger fish high in
the water column more readily. Such sources of variability should be kept
to a minimum.
Where applicable, those quality assurance/quality control (QA/QC)
procedures specified for the benthos should be followed for the coral and
fishes. These procedures include the maintenance of a voucher collection,
verification of identifications, inclusion of raw data sheets in reports,
etc. In addition, it is important that the applicant maintain consistency
between sampling periods in order to reduce sampling variability.
The monitoring report should contain a tabulation of the percentage of
living coral, coralline algae, and coral rubble for each photographic
quadrat. A tabulation of the number of individuals of each species of fish
identified on each transect should also be included.
A field notebook, as discussed in the benthos section above, should be
maintained and submitted along with the monitoring report.
Intertidal Communities—
Due to the offshore location of most marine sewage outfalls, monitoring
of intertidal communities will usually not be specified as part of the
biological monitoring program. Nevertheless, intertidal communities are
sensitive assemblages of organisms which may be affected by sewage
discharges (Dawson 1959, 1965). In cases where shoreward transport of the
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waste field is predicted, monitoring the intertidal community may be
required.
The type of intertidal community is largely determined by the type of
substrate (tidal flats, sandy beaches, rocky shores, gravel and cobble
shores), and the appropriate sampling procedures vary considerably among the
various communities represented in these environments. Conor and Kemp
(1978) provide a comprehensive review of the procedures used in quantitative
ecological assessments in the various types of intertidal environments; it
is unnecessary to provide a detailed reiteration of the sampling procedures
described therein for each habitat and biotic group. Selected sampling
procedures for rocky intertidal habitats will instead be described in order
to illustrate some of the basic principles involved.
A common attribute of many intertidal communities is the stratification
of the community with respect to tidal height, since many intertidal species
have discrete vertical limits within the tidal range. Since both community
species composition and density vary considerably with slight changes of
vertical distance (on the order of tenths of a meter), a sampling plan
designed to monitor the entire intertidal community should be stratified by
vertical height. Comparison of communities between reference areas and
discharge impact areas should be between samples from similar elevation,
also. If data for an entire transect across the intertidal community are
pooled, it is unlikely that differences between reference and discharge
impact areas could be detected because the within-transect variance produced
by combining data from different levels would be enormous. For an example
of the application of these principles, see Batzli (1969).
In order to detect differences between areas, or changes at one
location through time which arise from anthropogenic perturbations, these
differences or changes should be distinguished from the natural spatial and
temporal variations at a given location. Field sampling should, therefore,
quantify the spatial and temporal heterogeneity of each site. Intertidal
sampling is typically conducted along a transect. Samples are usually
collected (or a census of the community is conducted) at various locations
along the transect. These locations may be spaced evenly along the transect
(systematic sampling) or randomly along the transect (random sampling). The
advantages and disadvantages of each are discussed by Cochran (1977).
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Considerations for the selection of the number of transects, the length and
width of the transects, and the number and size of the sampling units are
discussed by Gonor and Kemp (1978). Samples are normally collected (or the
census of the community is conducted) within quadrats along the transects
(i.e., plots of constant area). The size of the quadrat is a function of
the nature of the species to be examined, their relative abundance, and the
cost of collecting (or conducting a census of) the organisms. Gonor and
Kemp (1978) recommend that a preliminary sampling program be conducted in
order to investigate the variability of a given area and to determine the
relative efficiency of various quadrat sizes and numbers of samples.
Sampling intertidal communities can be either destructive (in which a
quantitative sample of the biota is removed for later analysis), or
nondestructive (the acquisition of similar data using methods which do not
disturb the communty). Aside from a desire to minimize the damage done to
an area, there exists a second reason for favoring nondestructive sampling,
i.e., that repeated sampling of an area could be biased by the effects of
previous sampling. It is known, for instance, that reduction in the
abundance of certain keystone species (Paine 1969) may alter the rest of the
community in such a manner that change due to other events may be
undetectable.
Nondestructive sampling can generally be used if the species to be
sampled are visible, measurable, and unaffected by the sampling procedure
used (Gonor and Kemp 1978). Both the macroalgae and the sessile macrofauna
on rocky shores may be sampled without removal. Individuals within a
quadrat can be counted and measured in situ for some suitable dimension
(e.g., percent cover) which can be converted into an estimate of biomass
(Gonor and Kemp 1978). Nondestructive sampling is inadequate for both small
and mobile species, however, so if the entire community is to be censused,
some destructive sampling must occur. For most impact studies, however, it
may be sufficient to only examine effects on the plants and animals which
can be sampled nondestructively. Nondestructive sampling in the field may
be supplemented with photography (Littler 1971), which is particularly
useful when time in the field is severely limited.
Littler and Murray (1975) investigated the biological effects of a
low-volume domestic sewage discharge on the intertidal community on San
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Clemente Island, using sampling techniques which illustrate well the
sampling design considerations described above. Their study could be used
as a model for monitoring programs to be conducted in other intertidal
communities.
Littler and Murray (1975) determined the distributions and abundances
of macro-organisms with reference to tidal height and distance from the
outfall. They utilized a photographimetric technique (Littler 1971) for
assessing standing stocks (i.e., frequency and cover) of species popu-
lations. Sampling was restricted to macro-epibiota which could be discerned
with the unaided eye in the field or in photographs. Photographs were taken
of ring quadrats 30 cm in diameter (providing 0.07-m2 stratified plots) at
1- or 2-m intervals along transects both in the outfall area and at randomly
selected points in control areas. The control areas were sufficiently
remote from the influence of the outfall and had morphometry similar to the
area in the immediate vicinity of the discharge.
Cover was determined (Littler and Murray 1975) from the photographs
using a point-intercept method. If species were observed within a quadrat
but were absent from the scores, they were assigned a cover value of 0.05
percent. In some cases, samples contained multi-layered algal canopies;
thus, total cover was in excess of 100 percent. In these cases, more than
one photograph had to be taken per quadrat to measure stratification. Field
notes were taken using a tape recorder, which then facilitated later
taxonomic analysis of the samples.
Vertical heights for each quadrat were measured from fixed reference
points using a sighting level, a stadia rod, and standard surveying
techniques. Relative tidal heights were referenced to the level of
mean-lower-low-water (MLLW).
Sampling in different seasons (Littler and Murray 1975) showed that
seasonal changes in standing stock were minor, especially in the
sewage-affected area, so data for all sampling periods were considered to be
representative and were grouped as either outfall or control samples. If
seasonal changes had been significant, comparisons of outfall and control
samples would probably have had to be restricted to within a given season.
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The intertidal communities were stratified by 0.15-m intervals of tidal
height. The distribution of species populations as a function of tidal
height was then compared between the outfall and control areas. The
community features of species diversity, stratification, and species
assemblages were analyzed using the cover data. Cluster analysis was
utilized to objectively determine natural assemblages or groupings of
organisms.
The techniques utilized by Littler and Murray (1975) could easily be
applied to monitoring programs in other rocky intertidal environments.
While certain of the principles involved may apply to other environments,
sampling techniques will probably differ, and the more complete discussion
of Conor and Kemp (1978) should be consulted.
Analytical Techniques
Introduction--
The evaluation of compliance with BIP maintenance requirements
necessitates the analysis of biological monitoring data including
comparisons of spatial and temporal variability in the composition and
structure of the benthic macrofaunal assemblages and other selected biotic
groups among monitoring stations. A wide variety of techniques are
available for such analysis of biological data collected as part of a 301(h)
monitoring program. Analyses may range from rather simple qualitative
(tabular or graphical) comparisons of species distributions to complex
multivariate tests of the relationships of community structure to
environmental variables. The following sections provide discussions of the
applicability of several analytical techniques; however, no single approach
is recommended for analyzing the monitoring program data. In most cases,
the optimal approach will be to utilize several techniques. Selection of
the number and kinds of analytical techniques employed in each case will
depend upon the magnitude of the monitoring program (e.g., number of
stations and sampling frequency), the type of data collected, and the
distributional characteristics of the data. As an example, a biomonitoring
program of moderate complexity may include the following analyses of the
benthic macrofauna:
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• Calculation of community parameters such as diversity and
species richness
• Graphical or tabular display of abundances of dominant
organisms or indicator species
• Parametric or nonparametric statistical analysis of total
organism density, densities of individual groups, species
number, and community parameters
t Numerical classification of species abundance data (e.g.,
dendrogram).
Described below are analytical procedures recommended for use by the
applicant in conducting the required comparisons. Emphasis has been placed
on demonstrating the applicability of each analytical approach to the type
of monitoring required. These descriptions are intended as concise
introductions to techniques, each of which is the subject of numerous texts
and technical papers. Many of the more important references are included
under each description. The reference lists are not exhaustive, but they do
provide a starting point for gaining access to the literature.
The formulation of plans to analyze results properly occurs during the
development phase of the sampling design. Some statement must be made
concerning the expectation of the type of data that will be developed and
how that information will be used to address the issue of discharge impacts.
If preliminary studies have identified substantial site-specific
information, sampling objectives should be defined in detail. As discussed
above in the section concerning sample frequency and replication, the
ability of selected analytical techniques to detect differences in target
parameters among monitoring stations must also be assessed.
Limitations are inherent in each of the analytical methods described
below. Therefore, the inappropriateness of the singular use of any
technique is stressed. Individual methods or statistical models may be of
considerable utility in summarizing biological parameters; however, because
important assumptions are specific to each method, a single analytical
technique cannot accurately guide the interpretation of the monitoring
program results.
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Simple Graphic Displays—
Although statistical analyses of monitoring data are generally
necessary for demonstrating compliance with the 301(h) biological criteria,
graphical presentation of results provides an additional, and generally very
useful, means of making comparisons among sampling sites. In many cases,
the response of a biological parameter may be so pronounced that an effect
is clearly evident in a graphical presentation and a detailed statistical
analysis would not be necessary. Graphical displays are also important in
presenting summaries of large amounts of data in a concise format.
The following types of graphical presentations are recommended for
display of biomonitoring program results:
1. Community parameters (e.g., abundance, species richness,
diversity) at sampling stations
2. Trellis diagrams or dendrograms of station similarities
3. Maps of faunal assemblages near discharge (generally in
cases with large numbers of sampling sites in a
heterogeneous environment).
Examples of appropriate graphical displays of biological data collected near
marine sewage discharges are included in Figure 2. A discussion of
techniques is included in Green (1979).
Parametric Techniques--
Parametric statistical techniques such as Student's t-tests and
analysis of variance (ANOVA) are recommended for comparing measures of
abundance and community structure among sampling stations. When it is
hypothesized that outfall effects are evidenced by measurable differences
between monitoring stations, these statistical models can be used to
distinguish outfall-related impacts from natural variability in community
structure.
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OS
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Figure 2. Examples of graphical displays of biological data
from a marine sewage discharge site.
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As the name implies, these tests involve null hypotheses concerning a
statistical parameter of the variable being measured (e.g., population
means). However, as such, they have specified assumptions concerning the
distributional characteristics of the sample data. For ANOVA, it is assumed
that the error terms of the variates in each sample are independent and
normally distributed and that the sample variances are not different.
Independence of error terms is primarily associated with adequacy of
experimental design. The remaining two assumptions can be tested following
data collection. If the data do not meet the assumptions, transformations
can sometimes be applied to correct deviations from the assumed
distributions. Discussions of transformations prior to ANOVA are found in
Sokal and Rohlf (1969), Downing (1979), and Green (1979).
In many cases, a single transformation can correct both non-normal and
heteroscedastic data. It is important to note, however, that deviations
from normality, especially in cases of large sample size, will generally not
influence the overall test results to the same degree as heteroscedasticity.
Correlation of variances with means is a frequently encountered problem in
samples of organism abundances. In many cases, such violations of the
variance assumption can be corrected by a logarithmic transformation.
ANOVA is used to test the hypothesis that there are no differences in
the biological observations made at different sampling stations. In
addition to evaluation of a single factor (e.g., stations), ANOVA models are
especially appropriate for evaluation of the importance of multiple factor
level effects (e.g., depth, times) on the mean value of the dependent
variable.
The t-test is statistically equivalent to ANOVA when only two samples
are being compared. The t-test is appropriate for such two-sample
comparisons; however, it should be emphasized that the test cannot be used
to evaluate multi-sample hypotheses by testing all possible sample pairs.
In such cases the probability of committing a Type I error is considerably
higher than the designated level for each t test.
Discussions of the applications of ANOVA to biological data are found
in Zar (1974), Sokal and Rohlf (1969), and Winer (1971).
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If a significant effect is indicated in an ANOVA, an a posteriori
multiple comparisons test should be used to identify where differences are
located among the group means. The most commonly used a posteriori
procedures are the Student-Newman-Keul s test (SNK), the least significant
difference test (LSD), and Scheffe's test. Dunnett's test should be used if
only the control mean is to be compared with all other group means rather
than all possible comparisons. The characteristics of multiple comparison
tests are described in Zar (1974).
Nonparametric Techniques--
If sample data do not meet the assumptions of parametric statistical
tests, analogous nonparametric tests may be employed in the analysis of
differences among stations. Nonparametric tests do not utilize null
hypotheses associated with statistical parameters and there are typically no
assumptions concerning the distribution of the variates. An additional
advantage of nonparametric tests is that they can be used to test ordinal or
nominal data in addition to numerical values.
Nonparametric tests have a lower power (i.e., 1-3) than the analogous
parametric procedures. For example, a nonparametric ANOVA has a
power-efficiency of 95.5 percent when compared with the F test. Thus,
nonparametric tests should not be applied if the sample data meet the
assumptions of parametric techniques. Nonparametric tests are also unable
to test interactive effects in the ANOVA model.
Examples of some common nonparametric tests and their applications are
shown in Table 8. A comprehensive discussion of nonparametric techniques is
provided in Siegel (1956) and Hollander and Wolfe (1973).
As for parametric ANOVA, an a posteriori test should be used following
determination of a significant overall effect in the nonparametric analog of
ANOVA. A multiple comparisons test for equal sample sizes analogous to the
SNK test described in Zar (1974) and a procedure for unequal sample size is
presented in Dunn (1964).
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TABLE 8. EXAMPLES OF SOME NONPARAMETRIC STATISTICAL TESTS
Test
Mann-Whitney U-test
Kruskal-Wallis one-way ANOVA
Friedman two-way ANOVA
X2 test (or G-test)
Test of whether two independent samples
are from same population (analogous to t
test)
Test of whether K independent samples
are from different populations
(analogous to F test)
Test null hypothesis that K matched
samples are from same population
Test of independence of frequencies in K
samples.
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Multivariate Techniques—
Multivariate numerical methods are used to reduce and order large
matrices of data. They effectively summarize trends or patterns in the data
that are not otherwise observed from visual examination or univariate
analyses, and have been used to explore the interrelationships between sets
of biological and concomitant physical-chemical observations. Their most
common ecological application involves a search for patterns in measured
biological variables which can be related to patterns in measured
physical-chemical parameters. The goal in these analyses is to explain the
effect of environmental variables on both community composition and
structure. Examples of multivariate methods are discriminant analysis,
multivariate ANOYA, ordination techniques, and numerical classification
analysis.
Most multivariate tests have distributional assumptions analogous to
the univariate case, the most important of which is equality of dispersion
matrices. It is assumed that the vari ance-covari ance matrices are
independent of group means and are not different among groups. However,
with increasing numbers of variables (p) in the multivariate case, the
chance of detecting a significant difference becomes relatively high since
there are 0.5 p (p + 1) variances and covariances. Heterogenous
variance-covariance matrices will result in an increase in the probability
of a Type I error, i.e., that a significant difference between groups will
be indicated when one does not actually exist. In general, the potential
for variance heterogeneity can be considerably reduced by use of equal
replication, large sample sizes, and relatively few variables.
The increased probability of a Type I error is especially important
when the overall power of multivariate tests is considered. As indicated by
Green (1980): "When formal multivariate tests are made, their power
(especially with many variables) is so great that a significant result is
probable." Thus, the investigator should consider carefully the intended
purpose(s) of multivariate analyses before they are applied. Specifically,
it should be determined if the objectives are to reduce large data sets into
a manageable format, to evaluate general relationships to environmental
variables, or to test null hypotheses that no differences exist among sites.
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Numerical classification methods are used to distinguish groups of
entities (e.g., sample sites) according to similarity of attributes (e.g.,
species). Similarity of group attributes may be expressed using a variety
of resemblance measures, including commonly-used similarity coefficients
such as Jaccard, Bray-Curtis, Canberra metric, and Euclidean distance.
Classification begins with the compilation of a matrix of similarity
coefficients (index scores) between all possible pairs of entities. One of
a variety of available clustering methods is then used to form graphical
associations among entities to display groups of entities with similar
attributes.
In most ecological applications of classification methods, sample
collection sites are designated as the entities, and the relationship among
sites is defined in terms of similarity of species occurrence. This
approach is referred to as a normal classification, as opposed to an inverse
classification in which species are selected as entities and their presence
or abundance at the sample sites serves as the attribute. Analysis of the
monitoring station data set using both normal and inverse classification
methods, and the subsequent examination of normal-inverse coincidences using
a two-way table are recommended.
A description of classification methodologies, the use of numerical
classification, and an introduction to the literature concerning this
analytical approach are presented in the EPA report, Application of
Numerical Classification in Ecological Investigations of Water Pollution
(Boesch 1977). Reviews of important classification strategies are given by
Clifford and Stephenson (1975), Williams (1971), Sneath and Sokol (1973),
and Goodall (1973); and examples of ecological applications can be found in
Hughes and Thomas (1971), Boesch (1973), and Grossman et al. (1974).
Discriminant analysis summarizes multivariate information by weighting
individual variables so as to maximize differences in groups of entities.
This method describes differences between relatively homogeneous
species-assemblages (defined, for example, in a numerical classification
analysis) and facilitates identification of environmental variables which
best separate these groups. An introduction to discriminant analysis is
provided by Cooley and tonnes (1971). Pertinent examples of the use of this
method in the analysis of ecological data include Walker et al. (1979) and
Green and Vascotto (1978).
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Ordination refers to several multivariate techniques which are used to
reduce the dimensionality of a data structure and to relate biological
characteristics to environmental factors. The dimensionality is reduced by
one of several methods which are designed to minimize the loss of
information resulting from the reduction. In its most basic form,
ordination may be used to group similar sites based on biological
characteristics and to provide a graphical representation of between-group
relationships. Ordination in this manner is analogous to the production of
a dendrogram using numerical classification. A discussion of the use of
reciprocal averaging ordination as a classification technique is presented
in Gulp and Davies (1980).
Principal component analysis (PCA) and factor analysis are techniques
whereby axes scores in a reduced dimensional space are examined for
relationships with abiotic variables. Variables displaying high
correlations with component scores are assumed to be responsible for group
separation based on biological characteristics.
The relative merits of alternative ordination methods are compared by
Gauch and Whittaker (1972), and examples of environmental applications are
presented in Smith and Greene (1976), Sprules (1977), Gulp and Davies
(1980), and Hughes and Thomas (1971).
Multivariate ANOVA (MANOVA) is analogous to univariate ANOVA, but
includes measurement of more than one biological variable for each of
several samples and all measured variables are tested simultaneously. The
corresponding multivariate analysis of two samples is Hotelling's T2 test.
The basic assumptions are essentially the same as for the univariate case
(i.e., normality and independence of error terms and homogeneity of
within-group variance-covariance matrices). Although multivariate tests for
variance heterogeneity are available, their application is not recommended
by Green (1979) since they are more sensitive to the variance assumption
than are the MANOVA tests. Transformations such as the logarithmic may be
used to correct variance heterogeneity in the multivariate case so that
relatively minor violations of the assumptions do not seriously affect the
test results (Marriott 1974).
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Biological Indices--
Indices have commonly been employed in impact assessment of biological
communities because large amounts of multivariate data (i.e., abundances of
individual species) can be reduced to a single number. Indices can be
useful in this respect, but definite problems and limitations are associated
with their use.
A primary problem is the failure of investigators to recognize the
underlying assumptions and mathematical relationships of an index. By
overlooking such considerations, an index may be selected, applied, and
interpreted without a basic understanding of the properties of the
biological community which are actually being measured. Comprehensive
reviews of the assumptions and uses of diversity indices are provided in
Green (1979), Pielou (1977), Sanders (1968), and Peet(1974).
An additional problem associated with indices is that they may be used
extensively at the exclusion of other analytical or comparative methods
which retain more of the available information. Indices may supplement
multivariate techniques or analyses of individual taxonomic groups (see
preceding sections), but field and laboratory studies have demonstrated that
indices can be insensitive to rather intense biological change (Godfrey
1978; Swartz et al., 1980; Smith et al., 1979). Moreover, factors such as
sample size, collection method, and time of year may have a profound
influence on the value of an index (e.g. Hughes 1978). Therefore,
standardization of sampling procedures is a prerequisite to conducting
comparisons among index values.
Most indices commonly used in applied ecological studies are
descriptions of community structure (e.g., species diversity, evenness, and
richness). Other indices (e.g., Infaunal Trophic Index) incorporate
additional descriptive characteristics for each species and provide for a
description of community function. Several commonly-used diversity indices
are listed in Table 9. Species richness (S) and Margalef's index (d)
emphasize the number of species, are relatively simple, may provide valuable
biological information concerning impact assessment, and are much less
ambiguous than the information theory indices of Brillouin (H) and
Shannon-Wiener (H1). H is the diversity value for a sample, while H1 is an
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TABLE 9. A LIST OF COMMONLY-USED INDICES OF DIVERSITY
Index
Symbol
Equation
Species Richness
Margalef
Shannon-Wiener
Brillouin
Simpson's Index
where:
S
d
H1
H
SI
S = number of species.
N = number of individuals.
\i = number of individuals in the
number of species
S-l/ln N
-In
1-1
N N
N!
N (N-l)
species.
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estimate for a random sample from a larger population. H1 is used more
frequently due to the complexity of calculating large factorials in H.
However, for large numbers of individuals (N), the approximation, N (In
N-l), may be used for In N!, and computer programs are available for exact
calculation of H (Stauffer and Reish 1980). Pielou (1966) has suggested
that H is generally more appropriate than H' since a truly random sample is
required for estimation of H1. Simpson's Index is a measure of dominance
which is determined primarily by a few of the most abundant species.
Another index available for impact assessment is the Infaunal Trophic
Index (ITI) developed by Word (1978). The ITI is calculated as:
ITI = 100 -
33.33
where:
n- is the number of individuals in trophic group i.
The ITI is based on the relative proportions of individuals in four trophic
groups classified according to feeding types: suspended detritus feeders
(I), surface detritus feeders (II), surface deposit feeders (III), and
sub-surface deposit feeders (IV). ITI values correlate well with degree of
organic enrichment, in that decreasing ITI values indicate increasing
abundances of deposit-feeding organisms. The ITI is currently applicable to
benthic macroinvertebrate communities at depths of 20 to 800 m (66 to 2,625
ft) in the Southern California Bight. Research is currently being conducted
to determine applicability of the index to the Puget Sound area.
Since different ecological qualities are measured by the different
indices, it is recommended that species abundance data be used to calculate
at least three indices for each study: species number (S), combined number
of species and evenness (H1 or H), and dominance (SI). ITI would be an
important additional parameter for studies in the Southern California Bight.
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Indicator Species--
Many species display characteristic distributional responses to
pollutant sources. Analysis of the occurrence of such species, referred to
as "indicator species," may form a valuable component of the analysis of
sewage discharge impacts. The primary value of the indicator species
concept is that it allows for a considerable reduction in analytical
complexity, since individual abundances of a few species are used to
evaluate response of the community as a whole.
Indicator species may be divided into two categories:
• Sensitive organisms that display severely reduced abundances
near pollutant sources
• Stress-tolerant or opportunistic species that display
greatly enhanced abundances near pollutant sources.
In cases of organic enrichment, the first category of indicator species is
generally composed of suspension-feeding organisms such as Ampelisca
spp. (Amphipoda) and Amphiodia spp. (Ophiuroidea). Reduced abundances of
these species may also indicate high sensitivity to toxic chemicals
contained in the effluent.
The second category of indicator species, those having high abundances
in polluted areas, has received more intensive study than the former group.
Such species may have a high tolerance to organic enrichment or toxic
chemicals in addition to an opportunistic life strategy (e.g., short
generation time and/or lack of larval dispersal). These attributes enable
them to exploit available resources in the absence of nontolerant
competitors or predators following habitat disruption or pollutant stress.
A list of some polychaetous annelids that have been associated with
pollutant sources is provided in Table 10.
Although most of the species in Table 10 have been observed in high
abundances in very polluted areas, indicator species may also be used to
detect areas of moderate pollutant stress or transitional regions between
polluted and normal areas. Word et al. (1977) characterized species
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TABLE 10. LIST OF SOME COMMON POLYCHAETES
THAT HAVE BEEN ASSOCIATED WITH MARINE
OR ESTUARINE POLLUTION
Capitella capltata
Polydora ligni
Streblospio benedicti
Scolelepis fuliqinosa
Schistomeringos rudolphi
Dorvillea articulata
Heteromastus filiformis
Mediomastus ambiseta
M_. cal iforniensis
Eteone longa
Qphiodromus spp.
Cirriformia tentaculata
Neanthes succinea
_N. caudata
Source: Pearson and Rosenberg (1978)
89
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indicative of polluted and transitional areas of the Southern California
Bight. The clam Parvilucina tenuisculpta. the annelids Tharyx spp. and
Mediomastus californiensis, and the ostracod Euphilomedes spp. are present
in low abundances in control areas and reach much higher abundances (based
on absolute abundances and proportion of total infauna) in areas of organic
enrichment.
The primary limitation of the indicator species concept is that it
should be used only with a full consideration of the normal distributional
patterns and environmental associations of the species. This is especially
true in estuarine environments where salinity fluctuations and high organic
inputs may result in natural elevated abundances of opportunistic species.
In the marine environment, areas of natural organic accumulation (e.g.,
submarine canyons, kelp beds) may also have high abundances of opportunistic
organisms.
Before using an indicator species as part of a 301(h) monitoring
program, several types of information should be developed:
• Natural abundances of species in control areas
• Response of species to environmental conditions other than
pollutant stress
• Observed response of species to pollutant sources in the
biogeographic zone.
The primary requirement is that species abundance be adequately
described for control conditions. Proper selection of control sites will
ensure that any observed differences in the abundances of indicator species
are due to the discharge in question and not to natural or other
anthropogenic stresses.
Data Reporting
The information presented as part of a biological monitoring program
should consist of three general kinds:
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t Methods
• Study results and summary of findings
• Data reports.
A discussion of study methods should be presented in each report, including
such aspects as station locations, sampling procedures, sampling processing,
subsampling, quality control, and analytical methods. Procedural details
should be provided unless a standardized technique is used, in which case a
reference should be included. Citations should also be provided for all
taxonomic references used for organism identifications.
The presentation of study results should include a general
characterization of the biological communities sampled. Emphasis should be
placed upon descriptions of both spatial and temporal trends in community
structure and function. Specific comparisons should be conducted for all
biological criteria contained in the 301(h) regulations (e.g., ZID boundary
vs. reference communities). Where statistical analyses are performed, the
report should include details of the results. For example, in case of
ANOVA, the entire ANOVA table should be presented, not just a statement
concerning the significance level of the F value. Biological variables
(e.g., species abundances, diversity, richness) should be presented in
graphical or tabular format as means and their 95 percent confidence
intervals (X +_ t(n_j)Sx) for each sampling station.
Each monitoring report should include copies of the field collection
logs and laboratory sample counting forms. The data provided should include
the actual numbers of each species counted in each sample and the calculated
area! or volumetric abundance of each taxon. Sufficient detail should be
provided to allow for verification of analyses conducted as part of the
monitoring program, or for reanalysis of the submitted data.
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CHAPTER V
QUALITY CONTROL
The U.S. EPA policy on quality assurance and control (EPA Administrator
memoranda dated May 30, 1979, and October 14, 1981) stipulates that every
monitoring and measurement project must have a written and approved quality
assurance project plan. This requirement applies to all environmental
monitoring and measurement efforts mandated or supported by EPA. A quality
assurance project plan will specify the policies, organization, objectives,
and functional activities designed to achieve data quality goals of
individual projects or continuing operations. The monitoring programs of
301(h) permittees are covered by this policy.
Several EPA publications are available from the Office of Monitoring
Systems and Quality Assurance, ORD, USEPA, Washington, D.C., 20460, on the
subject of quality assurance. EPA Publication QAMS-005/80, for example,
describes 16 elements which should be included in all quality assurance
project plans. That publication establishes criteria for plan preparation,
including procedures to be used to document and report precision, accuracy,
and completeness of environmental measurements. In addition, the following
paragraphs provide guidance on quality control procedures specific to 301(h)
monitoring programs. The guidance provided below focuses primarily on water
quality monitoring and toxics control monitoring. Additional guidance on
quality assurance and control procedures is provided in Chapter IV,
Biological Monitoring, particularly the subsection on benthos.
APPROACH AND RATIONALE
In 301(h) monitoring programs, differences, or the lack of differences,
among samples must be demonstrated. When the differences to be defined are
small relative to background concentrations (as is the case for many
parameters), it becomes imperative to know and control the uncertainty
associated with sampling and analytical procedures. This is necessary to
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distinguish true field variability from that induced by sampling or
laboratory procedures.
It is imperative to have qualified personnel who are conscientious and
properly trained. Such personnel, using adequate equipment and procedures,
will help to ensure a clear understanding of sampling and analytical
variability relative to true field variability. A well defined quality
assurance/quality control program should be designed as an integral part of
the 301(h) monitoring program. The quality assurance/quality control
program should have as its basis a simple, but rigorous, quantitative
approach which can be applied consistently for the control of error. The
sampling and analytical procedures selected should provide feedback so that
those performing the analyses can promptly detect and correct procedural
problems. Thereby, the uncertainty regarding field measurement variability
will be minimized.
The section which follows describes quality control procedures
associated with field activities. A quantitative error analysis procedure
is then presented in detail with graphical techniques for detection of
increases in analytical error over time. Finally, quality control
procedures specific for toxic pollutant analyses are briefly described.
FIELD ACTIVITIES
It is important that field activities be well planned in advance and
that as many decisions as possible be made before field sampling commences.
Problems are encountered in coastal work which do not occur normally during
the sampling of inland waters. Highly conductive salt mists frequently
cause problems with electronic instrumentation. There is generally a lack
of nearby fixed points (landmarks, permanent buoys) from which to locate
sampling stations. Waves and swells niake the sampling vessel unsteady,
cause motion sickness, and make work difficult.
Field activity quality control can be subdivided into three categories:
1) accurate station location, 2) proper sampling procedures, and 3) proper
documentation of sampling efforts. Station location can best be assured by
use of redundant navigation systems. For example, the use of portable
electronic range positioning systems plus the use of a sextant in the
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horizontal mode would form such a redundant system. The two separate
procedures together help to minimize error. The installation onshore of
line of sight targets can greatly aid the positioning effort. Sampling
crews can line up the vessel from these first and then go to the position
fixing techniques for improved resolution. Finally, following sampling at a
station, the bottom should be sounded using a manual (non-electronic)
system, such as a lead line. The known depth (from charts and prior
measurements) and the lead line results can be compared for a rough check on
position.
There are two general types of sampling: 1) in situ measurements, and
2) the collection of water, sediment, and biological samples for subsequent
analysis. In situ measurements are normally made with electronic systems
(in situ observations by diver-biologists are discussed in Chapter IV,
Biological Monitoring). Calibration of these systems should be done before
and after each series of field measurements. Probe systems, except perhaps
for the hydrogen-ion electrode, are often unreliable in terms of absolute
accuracy and are best used in a differential sense (e.g., to measure changes
in parameters with depth while profiling at a given station).
Errors in sampling depth are caused by several factors including ship
motion and drag on the underwater sampling equipment. In high currents the
drag on instrumentation and cables may result in significant errors if
sampling depth is determined solely by the length of cable underwater. Drag
can be mathematically corrected for only if the current profile and the drag
coefficients for the instrumentation are known. The displacement of the
sampling device(s) from the desired position is calculated by either summing
the moments about the point where the cable enters the water or by making a
free body analysis of the forces on the sampling device and cable.
Many multi-probe in situ measurement systems incorporate depth
measurement by use of pressure transducers. The accuracy and precision of
such systems must be periodically checked. This is most easily done in calm
slack water. Precision is determined by multiple measurements at the same
depth. Accuracy is evaluated by comparison to measurements made with a
heavily weighted lead line.
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Dissolved oxygen probe systems should be calibrated using the modified
Winkler titration technique. Profile measurements of salinity and
temperature are used to determine water column stability and are an aid in
the prediction of plume behavior. Salinity probe systems offer moderate
accuracy but should be cross-checked by discrete water samples analyzed by
induction type laboratory salinometers. Temperature probes may, at best, be
accurate to within one-tenth to three-tenths of a degree Celsius.
Temperature probe systems are rarely linear over large temperature ranges
and must be checked against research grade laboratory thermometers.
Water quality sample collection, preservation, and storage should be
performed in accordance with the procedures discussed in Chapter III.
Procedures for taking biological samples are presented in Chapter IV of this
document.
An important aspect of receiving water sampling is the order in which
procedures are executed upon1 occupying a station. Vessel positioning should
first be completed. (Vessel location should be checked frequently.) Surface
observations should be made and recorded on standard sampling sheets (see
for example Figure 3). Water column samples are then collected. Following
that, the water column is profiled using in situ measurement techniques.
Next, any benthic samples should be taken. Only then should the depth be
sounded using a lead line or equivalent physical technique. Finally,
documentation should be completed before proceeding to the next station.
The major purpose of the above sequencing of activities is to prevent
detritus, resuspended during bottom sampling or depth sounding, from
contaminating water column samples or in situ measurements.
Documentation should include surface observations, sample log sheets,
and data sheets with results from in situ sampling. Waterproof preprinted
forms bound as log books have proven to be useful.
It should be reemphasized that the sampling program design must be well
detailed and that as many decisions as possible should be made before the
sampling crew starts their efforts. In coastal and estuarine sampling,
unpredictable problems will occur and will demand the immediate attention of
the sampling crew. Time for this type of on-site decision making must be
available.
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FIELD SAMPLING LOG SURFACE OBSERVATIONS
OJte (e.g. 09 June 1976)
enter; d.i^ (? digits), enter month (3 letters), enter year (4 digits)
Tide Stdorc pronounced. Ctests hive
• qldisy «(>|'(.'«r«nce 'OKl">»(wtlCiy I<0-^SO yiiTilst
500-l.tKKJ (wtrrt («|>|>rujti'n4te1r iSO yjtdv-5/8 (t,n )
1^ km (*(.|iru.t»4tfly S/8-I n m )
Z-* km («t'l'Mi»(-i«-ly I-? B * |
4-10 U (dj|>|>ri>«iiMtc-ly 7-t f> • )
10-7U t« (*(.("-fi-i«.*tolx 6-1? n M )
20-SD -- |JP|"<'. U,trl, 17OO n • }
SO km ur mure (30 * " ur iwirf)
rmsm wtAintu
(no iloud at any lt«
y [luudy (Kattrrrtt o
7 SPKIW. or rain and snow Hli
8 SlMwerfO
» IhwnihTSlomfil
l)
biok«n)
•High Water Sljck and Low Water Slack as measured at the southern point
of Port Jefferson Harbor. at the Bayville Brid'jc1 (Oyster Hoy), and at
tlic Lloyd Harbor entrance (lluntington Bay Complex).
Figure 3. Oceanographic surface observations log sheet.
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QUANTITATIVE ERROR ANALYSIS
This section presents techniques for error analysis. The focus is on
those parameters for which discrete water samples are taken in the field and
analytical work is performed in the laboratory. The error analysis
procedures are also applicable to in situ measurement. Analytical precision
is then determined by multiple measurements made over a short time period on
a discrete water sample, a procedure analogous to the splitting of samples
described in the following section. Accuracy can be determined by field
(e.g., probe) and laboratory analysis of the same sample, and the results
analyzed in the fashion described below for spiked samples (with the
concentration added set to zero).
Quantitative error analysis is based on the premise that every
measurement is subject to uncertainty. Uncertainty in water quality
measurements arises from the systematic bias and limited precision of
sampling and analytical procedures and from heterogeneity in the water
column at sampling sites. The quantitative determination of this
uncertainty serves as the basis of two essential steps in the control and
assurance of measurement quality:
t The routine monitoring of analytical precision and accuracy
• The presentation of results in a way that informs the
reviewers of the uncertainty of measurements and the
confidence which may be placed in conclusions drawn from the
results.
The first step in the quantitative analysis of errors is the
verification of a discharger's ability to produce analytical results which
are sufficiently accurate and precise to meet the specifications stated in
the table of recommended analytical methods in Chapter III. This
determination is made prior to the initiation of routine monitoring for each
water quality parameter to be measured. The recommended procedures for
determining precision and systematic bias are summarized in Table 11.
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TABLE 11. RECOMMENDED PROCEDURES FOR DETERMINATION OF SYSTEMATIC
BIAS AND PRECISION IN ANALYTICAL METHODS
Precision
Systematic Bias
• The precision is determined at four concentrations: one
concentration near the limit of detection of the procedure,
two Intermediate concentrations, and one concentration near
the upper limit of application of the method.
• Seven subsamples are analyzed at each of the four concentra-
tions.
• For procedures using analytical Instruments:
1) conduct analyses over a two-hour period.
2) run samples In the sequence high, low,
Intermediate, intermediate. Repeat
sequence seven times.
t Report the mean, standard deviations (Sa) and number of
samples analyzed at each concentration.
Add known amounts of the analyte to the low and one of the Intermediate
concentration samples used to determine precision. Enough additional
material Is added to double the lower concentration and to raise the
intermediate concentration to 75 percent of the analytical limit.
Seven subsamples of each spiked sample are analyzed.
Systematic bias Is reported as percent recovery at the final concentra-
tion of the spiked sample using the mean of the seven analyses.
oo
It 1s strongly recommended that actual samples be used In these analyses 1n order to Include the effects of naturally occurring Interferences.
These procedures can be adapted to nearly all the analytical methods specified In the monitoring program, except those using gas chromatography -
mass spectometry.
These procedures for determination of precision and systematic bias are taken from U.S. EPA "Handbook for Analytical Quality Control 1n Water and
Wastewater Laboratories" (1972).
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The initial verification of analytical competence is followed by the
continuous monitoring of analytical quality by each discharger or their
associated laboratory. It is essential that an unsatisfactory analytical
procedure be quickly identified and corrected in order to prevent the
accumulation of inaccurate data. This is accomplished primarily through the
routine analysis of split and spiked samples and the use of quality control
charts (as shown in Figure 4).
Laboratory analytical precision should be checked by splitting a
percentage of homogeneous field samples into replicates and analyzing all
subsamples. Sample splits should be made in the field. Laboratory
personnel should be kept unaware of which samples have been split. Ideally,
splits should not be analyzed in succession. For each split sample, the
individual measurements and the range are reported for each parameter. If
possible, a double blind procedure should be used.
The accuracy, which includes precision and systematic bias, of an
analytical procedure is routinely monitored by spiking a field sample with a
known amount of the analyte (a standard addition). As with the split
samples, personnel performing the analyses ideally should remain unaware of
which samples have been spiked. The deviation from stoichiometric behavior
of spiked samples is calculated from the following expression:
E = Xs - (xo + A) (1)
where:
E = deviation from stoichiometric behavior
Xs = measured concentration of spiked samples
XQ = measured concentration prior to addition of spike
A = increase in concentration due to spike.
The deviation from stoichiometric behavior of each spiked sample should be
reported along with the spiked and unspiked sample concentrations.
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-------
§
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SAMPLE NUMBER
SPECIFIED STANDARD DEVIATION - 0.1 mg/1
NOTE:
This figure illustrates how a decrease In measurement precision would affect
the distribution of data plotted on a precision control chart. In actual
practice, as soon as the loss of precision is detected—by about sample 62
in this case—the cause of the decreased performance can be isolated and
corrected. (Precision control data were generated using a mean oxygen
concentration of 5.0 mg/1.)
Figure 4. Example Precision Control Chart.
100
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The frequency with which split and spiked samples are analyzed must
represent a balance between reducing the likelihood of generating inaccurate
data and increasing laboratory costs for analyzing additional samples. The
U.S. EPA recommends that one split and one spiked sample be analyzed for
every 10 field samples analyzed (U.S. EPA 1979b). Those applicants who wish
to do so would have to justify a lower frequency of split samples. This
requirement can be applied to most parameters measured daily or weekly. For
parameters which are measured only at longer intervals (e.g., every month or
more), at least one split and one spiked sample should be analyzed during
each sampling period.
Project personnel other than the laboratory staff should be careful to
distribute the additional analyses over time and among sampling locations.
An even distribution of quality control effort over time permits continuous
monitoring of laboratory performance and ensures greater confidence in the
analytical results. Distributing quality control efforts among sampling
locations ensures that interferences present only at a few locations will be
detected and enables the precision in measurement to be determined for
different locations. It is especially important to be able to distinguish
analytical precision associated with effluent and receiving water sample
analyses.
The analysis of quality control data is facilitated by the use of
quality control charts. Individual measurements of analytical measurement
range and deviations from stoichiometric behavior are plotted on a graph
marked with the expected limits of deviation from the mean. Standard
practice is to draw "control limit" lines at three standard deviations from
the mean and "warning limit" lines at two standard deviations from the mean.
Since these correspond to 99.7 and 95.5 percent confidence limits, the
analytical procedure is considered out of control or potentially out of
control when the limits are exceeded more frequently than one in 370 and one
in 20 samples, respectively. In some cases, a developing analytical problem
appears as a diverging trend on a control chart even before the control
limits are exceeded. Examples of control charts illustrating cases when the
analytical procedure is under control and when it is not are given in
Figures 4 and 5.
101
-------
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SAMPLE NUMBER
SPECIFIED STANDARD DEVIATION * 0.1 w9/l
NOTE:
This figure illustrates how the presence of a systematic bias would
affect the distribution of data plotted on an accuracy control chart.
In actual practice,as soon as the systematic bias was detected—
before sample 60 in this case—the cause of the decreased performance
would be isolated and corrected. (Accuracy control data were generated
using a mean cadmium concentration of 0.8 ijg/1 [analytical standard
deviation - .1 pg/l] with 1.0 yg/1 added as spikes).
Figure 5. Example Accuracy Control Chart.
102
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The construction of quality control charts is a simple process. The
expressions necessary to calculate the limit lines are summarized in Table
12. Note that in many cases where the analytical variance shows a
concentration dependence, the quality control parameters can be adjusted to
account for this dependence. If most measurements of a parameter remain
within a small range (e.g., +_ 15 percent of the mean value), the assumption
of constant variance is sufficiently accurate for control chart purposes
even when the variance is concentration dependent.
For the purposes of verifying compliance with the precision and
accuracy specifications, quality control data should be plotted on control
charts derived from these same specifications. If a laboratory's precision
or accuracy is substantially better than required, additional limit lines
corresponding to the observed performance could be added to assist in
monitoring analytical performance more closely. Methods for generating
control charts from laboratory data are described in U.S. EPA (1979b) and
American Society for Testing and Materials (ASTM) (1951).
Analysis of unknown standards for all parameters should be performed at
least once a year in order to identify systematic error not detectable by
spiked samples. It is recommended that the dischargers or their selected
laboratories participate in the U.S. EPA interlaboratory testing program.
The interlaboratory correlation technique of Youden (1960) could be used for
comparison of results.
As a final check, all submitted data should be critically reviewed. It
may be requested that unusually high or low measurements be reanalyzed.
The second function of quantitative error analysis is to facilitate the
presentation of data in a way that permits open inspection of their
certainty or uncertainty. This requires quantitative estimates of both the
variability introduced in measurements by analytical imprecision and field
heterogeneity and the limits of sensitivity of analytical methods.
The variability inherent in the analytical procedure is determined from
the results of split sample analysis. If the variance is relatively
constant over the range of measured concentrations, a pooled estimate of the
analytical variance can be made using the following expression:
103
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TABLE 12. SUMMARY OF EXPRESSIONS NECESSARY
TO CONSTRUCT CONTROL CHARTS
z
o
t/>
(_>
UJ
IX
0.
£1
>-
(_>
<_>
-------
1 NS
S a " 2N ^ i
s i = l
where:
o
S g = analytical variance
NS = number of split samples pooled
R^ = range of subsample measurements for sample i split n ways.
If the variance is concentration dependent, a different parameter, e.g., the
coefficient of variation, can be averaged.
The variability in measurement caused by field heterogeneity is
quantitatively determined by the analysis of replicate field samples. Two
sampling strategies should be considered:
1. Collect replicate field samples and analyze multiple
subsamples of each. Analysis of Variance (ANOVA) is then
applied to determine the contribution of field heterogeneity
to the variance. The drawback of this approach is that it
requires a large number of analyses for even minimum
resolution power, e.g., 4 x 4 = 16.
2. Collect replicate field samples and analyze each only once.
The field-induced variance is estimated using the principle
of variance additivity. This approach makes additional
assumptions concerning the analytical variance but requires
many fewer analyses to be made, e.g., 4 x 1 * 4.
Since the application of the first approach is well understood [see for
example Sokal and Rohlf (1969)], only the second will be described in detail
here.
105
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The second approach is based on the assumption that an independent
estimate of the analytical variance exists which is applicable to the
conditions under which the field replicates are analyzed. If this is the
case, the field variability can be estimated from the following expression
[American Chemical Society (1980)]:
where:
S2f = variance due to field heterogeneity
S2 = variance of replicate field samples
S2a = analytical variance.
a
Equation 3 can be applied using the pooled estimate of analytical variance
(Equation 2), if one the following conditions is met:
1. The analytical variance is not strongly dependent on analyte
concentration or background interferences.
2. The concentration of analyte in samples used to compute the
pooled or initial analytical variance is similar to that in
the field replicates.
Replicate sampling should be conducted at all field stations where
measurements are to be used in comparisons. Analysis of replicate sample
data is necessary for assessing the reliability of such comparisons. When
replicate sampling at all stations is not feasible, replication is required
for at least one station from each group of stations which may reasonably be
assumed to have similar amounts of field heterogeneity. As a minimum,
replicates should be collected at one ZID boundary station and one reference
(control) station.
The number of replicates to be collected at each station depends on the
use intended for the data. For example, more replicates are required to
make a meaningful comparison of variances than are required to compare mean
106
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values. Appropriate statistical methods should be applied to each case.
Sevenfold replication is currently recommended by U.S. EPA (American
Chemical Society 1980). Depending upon circumstances, overall project
design, and the applicant's resources, fewer replicate determinations may be
accepted.
If possible, replicate samples should be taken at all specified
stations within the first year of the sampling program during a period of
maximi/m natural variability. This would provide information on field
heterogeneity to be utilized during the sampling program in-progress review
(review of first annual report). At that time, the replication program
should be evaluated and re-designed if necessary. In addition, the
knowledge of field variability can be applied to design a composite sampling
scheme for all stations. By analyzing one sample formed by combining and
homogenizing a number of replicate samples, the uncertainty in the
measurement due to field heterogeneity can be reduced without performing any
additional analyses. The following expression defines the number of
replicates required to obtain a confidence limit of E for a mean value:
Nr = (t • Sf/E)2 (4)
where:
t = appropriate value of Student's distribution
Nr = number of replicates required
Sf = field variability: Sf » Sa.
When the condition of negligible analytical variance is not met, composite
sampling is of little value.
In addition to the analytical and field variances, knowledge of the
limits of quantisation and detection is essential to the assessment of
measurements at trace levels. The method recommended by the American
Chemical Society (1980) is based on the analytical variance determined using
a field blank as follows:
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L = 3 • S
LQ = 10 - S
where:
LD = limit of detection
LQ = limit of quantisation
SB = ^St + sb wnere st and sb are the standard deviations
of an instrument response to replicate instrument runs of a
single analyte - containing sample and a blank sample,
respectively.
For a more detailed analysis of limits of detection and quantisation, see
Currie (1968).
TOXIC POLLUTANT ANALYSIS
Assuring the quality of toxic pollutant analyses requires numerous
precautions beyond those necessary for other water quality parameters. Most
of the additional quality control procedures are necessitated by the greater
complexity of analytical instruments used for toxic pollutant analyses and
the risks of sample contamination. Requiring all toxics analyses to be
conducted by U.S. EPA certified laboratories helps assure the adequacy of
internal quality control practices. Each laboratory should be practicing
the quality assurance steps outlined in the source documents for each
procedure as well as appropriate methods in the quantitative error analysis
portion of this chapter. Special quality control procedures for toxics not
dealt with in the above sources are the concern of the remainder of this
section.
Extra care in sample handling is required for toxic pollutant analysis
since these pollutants generally occur in trace concentrations and
frequently are unstable. Samples should be stored in the dark to avoid
photochemical decomposition. Storage at reduced temperatures, as specified
in Chapter III, minimizes the rate of other degradative chemical reactions.
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Exposure of the sample to the atmosphere should be minimized in order to
avoid loss of volatile compounds.
Sample bottles must be clean and made of materials which will not
contaminate the samples. Plastic or glass bottles must be specified
depending upon the analyses to be performed on the sample. Caps for all
glass bottles used to store toxics samples should be teflon lined. Glass
bottle and cap liners should be cleaned with chromate cleaning solution and
successively rinsed with distilled water and several portions of the
appropriate spectral grade redistilled organic solvent. Bottle caps should
be washed with detergent and rinsed using the same steps described above.
Plastic sample bottles should be cleaned with detergent or concentrated
hydrochloric acid and rinsed with distilled water.
Since many organic substances are strongly sorbed by particulate
matter, it is essential that effluent samples contain a fraction of
suspended solids representative of the entire waste stream. This should be
considered in the selection of sampling devices and locations. The
variability in measurements introduced by sampling techniques, together with
the variability caused by effluent heterogeneity, are considered below.
The assessment of the significance of toxic pollutant measurements
depends on a knowledge of the variability in measurements introduced by
sampling technique and site heterogenity. This variability is determined by
the analysis of replicate samples. The rationale used to specify the number
and frequency of replicate samples is contained in the Quantitative Error
Analysis portion of this chapter. Replicate samples should be collected
from the effluent stream and in most cases from the sediments of at least
one field station. Replication should be carried out at least once a year
during the period of highest toxic pollutant levels.
The qualitative and quantitative analytical capabilities of a
laboratory should be verified as well. One effluent sample spiked with each
of the routinely monitored substances (except dioxin) should be analyzed
with each group of samples. A blank (glass doubly-distilled water) should
be analyzed along with each group of samples screened for priority
pollutants.
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APPENDIX A
OCEANOGRAPHIC METHODS
This appendix is intended to provide background information and
guidance on collection of oceanographic data. Types of current meters and
their proper use are summarized. The use of drogues, drifters, and dye
studies are also reviewed. In addition, specifications for field use of
current meters, drogues, drifters, and dye studies are discussed.
Positioning methods are briefly reviewed.
CURRENT METERS
Uses of Current Meters
Current meters are used to measure the variation in current speed and
direction at a fixed location with time. Since each meter collects data
only at a single point, several current meters may be required to establish
the velocity field over depth and over a given area. A typical fixed
current meter array may contain a meter near the bottom, one near the
surface, and one at mid-depth. Short-term current measurements can be taken
from a boat with a meter lowered over the side and connected by cable to an
on-deck recorder. Longer term in-situ measurements may be obtained by
installing meters with self-contained recording devices (e.g., magnetic
tape, film, and strip charts) on a mooring system anchored to the bottom.
Current speed and direction may be recorded either continuously, or at
specified time intervals. Vector averaging meters electronically average
the measured velocity components over specified sampling intervals, and
provide the average current over each interval as output.
Different current meters are designed to work in different operating
environments. Meters designed to operate at greater depths must be able to
withstand higher pressures, and must be capable of accurately measuring
relatively low current speeds. Meters designed to operate in shallow
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nearshore coastal environments must be able to accurately measure currents
in the presence of wave action. The effects of waves include both
oscillatory water paYticle orbits in the immediate vicinity of the meter,
and the pumping action of the mooring lines where a surface or subsurface
buoy is subject to wave-induced motions.
Types of Current Meters
The current meters presently in use may be classified as mechanical,
electromagnetic, or acoustic. Mechanical current meters include Savonious
rotors, ducted impellers, drag inclinometers, and propeller-type meters.
Savonious rotor current meters utilize a unidirectional rotor on a vertical
axis of rotation and a vane which senses the horizontal direction of flow.
Although these meters have been used for many years in oceanographic work,
they are not suited for operation in shallow water environments which are
exposed to wave and swell activity. In such waters the oscillatory velocity
components due to the orbital motions of waves are significant. Since the
rotor turns in only one direction, the oscillatory velocity components due
to waves are recorded as a rectified velocity input. The rectified record
cannot be adequately resolved into the oscillatory components since the
directional response of the meter is slow relative to the wave motions, and
its response to deceleration is different than its response to acceleration
(Horrer 1968). As a result, the velocity record is distorted and never
reaches zero velocity, even in the presence of wave action alone. This
makes it difficult to distinguish the steady component of the current from
the oscillatory component due to waves. Savonious rotors should, therefore,
only be used in deep waters below the influence of wave action or in areas
where the current speeds are high and the waves very small.
Two types of ducted impeller current meters are available which attempt
to eliminate the problems associated with current measurements in the
presence of waves. One type manufactured by Endeco is a neutrally buoyant
meter with a bidirectional impeller which is attached to the mooring line by
a tether several feet in length. The tether and neutral buoyancy allow the
meter to move with wave-induced water particle orbits, so these short-term
velocity oscillations are not recorded. The meter is said to orient itself
in a wave field so that only the mean current is measured. A tilt
compensation mechanism keeps the meter horizontal (Brainard and Lukens
111
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1975). Another horizontal ducted impeller meter manufactured by Bendix has
a long boom vane to keep the meter oriented in the direction of mean flow,
preventing it from rotating in the presence of short-term wave-induced
oscillations (Horrer 1968). The ducted impeller is bidirectional, and the
oscillatory components of the horizontal current velocity are recorded but
can be separated from the record.
Davis-Weilar propeller type meters are also designed to operate in the
presence of waves. These meters utilize two orthogonal propeller assemblies
designed to respond linearly to the current vector components so that
averaging or filtering of oscillatory movements can be done properly (Wald
1979). Electronic circuits resolve the signals from the propeller sensors
into N-S and E-W velocity components, and electronically integrate them over
a specified sampling interval. This gives the vector averaged-current over
the interval. Oscillatory movements associated with wave action are removed
by electronic filters in the averaging procedure.
Drag inclinometer type meters consist of a cylinder with stabilizing
fins which is suspended from a pivot point at one end. The drag and lift
forces due to the current velocity deflect the meter at a vertical angle
which is measured by an inclinometer. The horizontal direction is measured
by a compass. Although this type of meter is claimed to measure currents in
the wave zone accurately, it may suffer some limitations under these
conditions since its response (tilt angle versus current speed) is not
linear. This nonlinearity makes it difficult to separate the oscillatory
velocity components from the mean current (Daubin et al., 1977). Also,
vertical wave-inducted motions will deflect the meter vertically, and
movement of the mooring line or significant turbulence in the flow may
result in measurement errors.
Electromagnetic current meters measure the instantaneous x- and
y-velocity components at a flow sensor which contains a wire coil and two
orthogonal pairs of electrodes. The coil produces a magnetic field, and the
electrodes measure the voltage gradient across the coil which is induced by
the water as it flows through the field. The orthogonal electrode pairs
measure the x- and y-velocity components. Since the voltage gradient is
proportional to the current speed and there are no moving parts in the
sensor, the meter is capable of a fast, linear, highly sensitive component
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response (McCullough 1977). These characteristics make the meter suitable
for use in the presence of waves, since the oscillatory components can be
separated either from the records or filtered out by electronically
integrating the electrical signals over a specified sampling interval.
Acoustic current meters determine current velocities by measuring the
relative travel times of two simultaneous acoustic signals transmitted
across the flow sensor. The transmitted acoustic beams are focused on a
reflecting plate which returns the signals to the receiving transducers. An
acoustic phase shift detection scheme correlates the travel times with the
current velocity component along the beam path. Two pairs of orthogonal
transducers measure both the x- and y-velocity components. Voltages
proportional to the velocity components are resolved electronically using
the output signals from the transducers and compass. Since acoustic flow
sensors have a fast, linear, highly sensitive response (McCullough 1977),
they are suitable for use in the presence of waves. The wave-induced
components may be separated either from the records or filtered out by
electronically integrating the records over a short sampling interval.
DROGUES AND DRIFTERS
Use of Drogues
Drogues are used to trace the path of moving water near the surface or
at fixed depths below the surface. The drogues are released at a given
station and are subsequently tracked by recording their positions at short
time intervals. The sequence of positions and travel times between
positions gives information on the actual path a particle may travel in the
currents, and the mean velocities between points along the path. However,
the actual paths traced by two drogues simultaneously released at a given
point will rarely be the same due to the random components of the velocity
field. Thus, several drogues should be released together.
Drogue studies can be used to determine the current patterns in the
vicinity of an outfall and to evaluate the movement of a waste plume which
either surfaces or remains submerged at a given depth. If a sufficient
number of drogues are released at appropriate depths and are monitored at
frequent intervals, information can be obtained on the net transport and
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horizontal dispersion of a waste field, and on the variations in the current
velocities along the path followed by the plume. If the drogues are
followed long enough, they may indicate where the plume will reach the
shoreline, the length of time involved, and the path followed between the
discharge point and the shoreline.
Types of Drogues
Most drogues can be classified into one of the following four
categories: 1) parachute drogues, 2) cruciform drogues, 3) window shade
drogues, and 4) cylindrical drogues. Of these types, parachute and
cruciform drogues are the most widely used. Parachute drogues consist of a
passenger parachute or smaller pilot parachute which is usually attached to
a weighted vertical pipe. The parachute is supposed to remain open and
oriented horizontally with its opening facing the currents. However, since
some parachutes are denser than water, they tend to hang downward in a
closed position at low current speeds. This problem can be reduced by using
a spreader bar or ring to help keep the parachute open, and by adding
buoyance to the parachute so that it becomes neutrally buoyant. Cruciform
drogues usually consist of two or sometimes three identical slotted sheets
of plywood, masonite, plastic, or metal arranged in a bi-planar (or
tri-planar) crossed vane. The vanes can also be constructed of canvas or
cloth stretched across a solid frame. Window shade drogues consist of a
rectangular sheet of plastic film, canvas, or cloth suspended from a
spreader bar and bridle at the top, with a weighted spreader bar at the
bottom. The rectangular drogue should hang approximately vertical with its
plane surface oriented perpendicular to the direction of the horizontal
flow. However, this orientation, which maximizes the drag forces of the
currents, may not always be achieved under actual field conditions.
Cylindrical drogues include various vertically oriented cylindrical objects
which have been used to trace currents (e.g., 55 gallon drums, drift poles,
wooden barrels).
Most drogues areyreally drogue-buoy systems consisting of a small
marker buoy which is tracked at the surface and a larger submerged drogue
portion with is set at the desired depth by a connecting line between the
two. The drogue portion must be weighted and ballasted so that the drogue
assembly has sufficent negative buoyancy to keep both the drogue and
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connecting line in their intended vertical orientation, and to keep the buoy
mast upright. The connecting line must remain close to vertical so the
drogue will be measuring currents at the desired depth.
Drogues are intended to passively drift with the currents at a
specified depth. In reality, some errors are introduced in the drogue
trajectories by wind drag on the exposed portion of the marker buoy, by the
relative surface current drag on the submerged portion of the surface buoy,
and, for deep drogues with long lines, by the relative current drag on the
connecting line. Since surface currents are generally faster than the
deeper currents, it is important to design drogues so that the projected
area of the submerged drogue is maximized and the projected area of the
surface buoy is minimized. Large drogues are much more difficult and
cumbersome to launch and retrieve from a boat than smaller designs,
especially when many drogues are involved. However, since the accuracy of
the measurements generally increases with the size of the submerged drogue,
it is best to use the largest size practicable in a given situation. The
surface buoy should be the minimum size required to keep the system buoyant
and trackable at the surface. The wind forces on the exposed portion can be
minimized by using a buoy which is almost completely submerged and which has
only a thin radio antenna or small radar transponder protruding upward for
tracking. Buoys with considerable freeboard or with larger tracking devices
such as flags, radar reflectors, or flashing lights may be subject to
significant external wind forces. If possible, it is best to avoid
performing drogue studies under high wind conditions, especially when trying
to measure the lower current speeds in deeper waters.
Uses of Surface Drifters
Surface drifters are used to measure the average path of currents at
the surface. Drift bottles and drift cards are the types most commonly
used. Vertical drift cards and drift bottles are useful for evaluating the
movement of effluents which reach the surface. Horizontal drift cards,
which float horizontally on the surface, may be used to determine the
potential movement of surface slicks due to oils or other floatables which
form a surface film. Because these movements are influenced largely by the
wind, horizontal drift cards should be released under several different
meteorologic conditions.
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Surface drifters provide a rough estimate of the travel times between
the release and recovery allowing a net drift rate to be computed. No
information is given on the actual flow path or the velocity variations
along the path. However, they do provide information as to whether or not a
surface waste field can be expected to reach the shoreline, and, if so,
where it will make contact and approximately how long it may take.
Types of Surface Drifters
Drift bottles are long-necked glass bottles partially filled with sand
ballast so that only 0.25 to 1 in of the bottle neck remains above the
surface. The bottle size is typically 4 to 6 ozs (Grace 1978). Each bottle
contains a readily visible postcard with instructions requesting the return
of the card with information on the location and time of recovery and
generally the offering of a reward.
Drift cards are available in several different forms. Most common
types are either drift envelopes, which are plastic envelopes containing
instructions on a return card, or plastic drift cards, which are rectangular
in shape with identification numbers and instructions stamped into them. As
with drift bottles, the information requested is the location and time of
recovery and a reward is generally offered for their return. Horizontal
drift cards are designed to float horizontally and, therefore, measure
transport in the surface film or upper millimeter of the water column. As a
result, they are strongly influenced by the wind and do not really measure
the surface currents. Vertical drift cards are designed so that only one
edge remains at the surface, with the card retaining a vertical orientation
in the water column. This is accomplished either by using a negatively
buoyant card with a foam flotation strip on the upper edge or a positively
buoyant card with a weight strip on the lower edge. Both vertical drift
cards and drift bottles measure the average horizontal transport in the
upper few feet of the water column, and thus provide a measure of the
surface currents. A good design will attempt to minimize the effects of
wind on the drifter by minimizing the ratio of the sail area (area exposed
to the wind) to drogue area (submerged area exposed to current). Drift
bottles may be better in this respect since the exposed bottle necks are
much narrower than the submerged part of the bottles.
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Since the recovery rate of surface drifters is generally low, many
drifters must be released at each station in order to obtain an adequate
amount of information. The drifters are usually deployed by boat, but may
also be released by airplane. Because of their compact size, it is easier
to release a large quantity of drift cards than drift bottles. A good
drifter must be durable enough to survive at sea, reach the shore through
the surf, and should attract attention once it reaches the shore.
Use of Seabed Drifters
Seabed drifters measure the average path of currents near the sea
floor. They are useful for determining the fate of waste materials subject
to transport by bottom currents. This includes settleable solids and any
portion of the effluent which remains near the bottom. The drifters provide
information on the net movement of a waste field along the bottom, including
where the waste field may reach the shoreline, and a rough estimate of how
long it may take. If a sufficient number of drifters are recovered, they
may indicate areas of possible shoreline contamination.
The success of a bottom drifter study depends on reasonable recovery of
the drifters. The drifters generally have labels attached which request the
location and time of recovery and promise a reward. The condition of the
drifter should also be recorded since, if the rod becomes detached, the
saucer will float and therefore measure the surface currents rather than the
bottom currents. The drifters may be recovered either offshore by
commercial fishermen using bottom trawls or bottom gill nets, or more often
at the shoreline. The recovery rate will depend on recreational access to
the beaches, the intensity with which the beaches are used, and on how
extensive the commercial fisheries are in the area. Recovery rates from
less than 5 percent to over 50 percent have been reported. Therefore, many
more drifters should be released than are expected to be recovered in order
to obtain a sufficient amount of data. The information obtained from
drifters recovered offshore by commercial fishermen is probably more
accurate than data from shore-recovered drifters, since in the latter case
no information is provided on the time of first contact with the shore or
possible movement in the surf zone. In either case, the only information
provided is the point of release, the point of recovery, and a rough
estimate of the travel time between the two points.
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Types of Seabed Drifters
Woodhead type drifters are generally used. These devices resemble
umbrellas in shape, consisting of small plastic dished saucers with a long
thin plastic rod attached at the center. Typical dimensions are 18 cm (7
in) for the diameter of the saucer and a 0.65 cm (0.25 in) diameter rod
about 54 cm (21 in) long. The rod terminates in a sharpened point and a
small, weighted collar [about 6 g (0.2 oz)] is attached near the end (Grace
1978). The saucer has a slight positive buoyancy so it tends to hover
slightly above the sea floor and drift with the bottom currents; the weight
collar causes the pointed end of the rod to lightly drag along the bottom.
Seabed drifters may be deployed either at the surface by a boat or
low-flying small plane, or they may be taken to the bottom and released by
divers. Several drifters may be released at the same location on the
bottom, even if deployed at the surface, by attaching them to a salt spool
which dissolves after the drifters reach the bottom. However, because the
rate of descent of the drifters is fairly slow, the actual release point on
the sea floor will differ from the known release point at the surface. This
difference increases with increasing depth and subsurface current
velocities.
DYE STUDIES
Dye studies, using fluorescent tracers, can be very useful in
determining the behavior of waste plumes, as well as indicating general
circulation patterns in the vicinity of discharge sites. A suitable tracer
can be injected either directly into the waste stream before it discharges,
or into the receiving water at some point near the discharge site. The
injection may be a single dose or a continuous release. The movement of the
waste plume and the horizontal and vertical dispersion of the effluent can
be determined by measuring the concentration distribution of the dye tracer
both temporally and spatially after the initial injection. The dilution
rates and the spatial and temporal distribution of contaminants at a given
distance from the discharge site can then be evaluated. Three-dimensional
distributions can be obtained by sampling at various depths.
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SPECIFICATIONS FOR FIELD WORK
After selection of the appropriate method(s) for obtaining the
necessary oceanographic information, specifications are needed for the
general types of equipment to be used, number of measuring devices, location
of measurement stations (drogue/drifter release points or dye injection
points), and frequency and times of measurements. If a dye study is to be
done, type of dye, appropriate concentration, measurement technique, and
frequency should be specified. This section discusses specifications for
current meters, drogues and drifters, and dye studies.
Current Meters
The type of fixed current meter selected depends on the importance of
wave motion at the site and available equipment. The wave motion expected
at the site should be evaluated to determine if data obtained could be
affected by wave-induced orbital motion. If so, a non-Savonius type of
curent meter should be used. Calibration and visual inspection should be
made of the current meters before and after use to detect other sources of
error such as fouling of the rotor or the sensor probes. In-situ current
meters set for at least 5 days at a time are preferred to current meters
deployed for short periods from a boat. The current meter array should
include meters set near the surface, at mid-depth, and 1.5 m (5 ft) above
the bottom. The exact depths are determined after reviewing the available
current data to locate depths where different currents exist, the expected
height of rise of the plume, and the current measurement objectives (e.g.,
whether movement of the waste field or sediment movement is of primary
interest).
Current meter location decisions also depend on the current measurement
objectives. One meter array should be located near the discharge site. As
a minimum, currents should be measured continuously or at least for 15 to 30
min during all 5-day spring and neap tide periods, on a quarterly basis,
when receiving water samples are collected.
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Drogues and Drifters
The resolution of information obtained during a drogue or drifter study
depends on the number of drogues or drifters followed, the frequency with
which positions are fixed (recorded), and the accuracy of the method used to
fix positions. The use of drogues is recommended over drifters because
drogues give information on the flow path and current speeds along the path
and also because of the low recoverability of drifters and poor estimates of
travel times obtained from drifters. However, these factors have to be
considered along with the cost of tracking the drogues.
If certain types of drogues are more suitable for the wind and current
conditions at the site, then those types should be recommended in the
monitoring program. The drogues should be released over the diffuser at the
approximate level to which the plume rises during the time of year of the
study. At least five to six drogues should be released each time. The
drogues' positions should be traced at half-hour intervals up to a distance
of 3.7 km (2 nmi) from the outfall or for a total of 10 hr. Successive
vector plots showing drogue position over time should be provided on
large-scale nautical charts.
Drifters could be used if the primary concern was whether a surfacing
waste plume might reach shore or if only an approximate travel time were
needed. The choice of drifter type would depend on deployment method, cost,
and distance from shore. Vertical drift cards are the easiest to use and
are more durable than glass bottles. The drifters should be released at the
point of discharge or the middle of the diffuser.
Dye Studies
Specifications for a fluorescent dye study to be done as part of a
given monitoring program include the time of year for the dye study, the
type of dye to be used, the approximate concentration to be achieved in the
receiving water, the length of time of dye injection and measurement, and
suggestions on the measurement technique to be used. Each of these aspects
will be discussed briefly.
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The purpose of a dye study is to determine the horizontal and vertical
extent of the plume and the direction of the plume's movement. A dye study
should be performed at a time of minimum stratification, which occurs most
often in the fall or winter, and at a time of maximum stratification, which
occurs most often in the summer. Density profiles collected during water
quality monitoring surveys should be checked to identify or verify periods
of maximum and minimum stratification. The selection of the type of dye
depends on the degree of adsorption which could occur at the site, the cost,
and availability of chemicals. The dye tracers most commonly used at the
present time are rhodamine B and rhodamine WT. Rhodamine B has fairly
strong adsorptive tendencies, and may be adsorbed onto suspended solids,
sediments, plankton, aquatic plants, and sampling and injection equipment.
Rhodamine WT is the preferred choice because it is much less susceptible to
adsorption than rhodamine B, although it is about twice as expensive. Both
dyes are available in liquid form, which is strongly recommended over
powdered forms due to ease of handling. The solutions should be adjusted to
the same density as sewage effluent, if necessary, by mixing with methanol
or by addition of salt.
The effects of temperature, degradation, and photodecomposition must
also be considered. The fluorescence of a sample will vary with
temperature, although this can be adjusted easily with a temperature
correction curve. Chemical degradation can be a problem in the presence of
strong oxidizing agents such as chlorine. The degradation rate is usually
low at the chlorine levels typically encountered in the field, but it should
be evaluated for dye injected directly into a chlorinated waste stream.
Photodecomposition rate can be estimated from a control solution of dye
which is exposed to sunlight on the boat and monitored for the duration of
the study.
Concentrated dye solutions must be diluted to appropriate levels before
being released. Because the fluorometer calibration curve reverses
direction for dye concentrations greater than about 1.0 ppm (Turner Designs
1976), dye studies should be designed so that the dye concentrations
measured in the field are below this level. Dilution techniques can be used
for higher levels where discrete water samples are collected. However, this
is not done easily when an on-ship flow-through apparatus is used or when
in-situ measurements are taken with a towed submersible fluorometer. The
121
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dye should be added to the effluent just before it enters the outfall for a
period of about 6 hr, begining 3 hr before the current reverses. The
concentration in the effluent stream should be 0.2 ppm above the background
fluorescence of the seawater. The background fluorescence of moderately
polluted water may be as high as 100 to 200 parts per trillion, and raw
sewage has a background fluorescence of about one part per billion (Turner
Designs 1976). The pigments present in blue-green algae can yield
substantial background fluorescence if large concentrations are present.
Although appropriate optical filters can be used to minimize this problem,
background fluorescence should always be determined from seawater samples
before releasing the dye.
The dye can be measured using a submersible fluorometer or a continuous
pump and shipboard fluorometer. Submersible fluorometers, if used, should
be towed at a depth approximately equal to the equilibrium level of the
plume. Measurements should be made for about 12 hr beginning 3 hr before
the start of the dye injection. When dye studies are conducted, turbidity,
temperature, and salinity profiles should be taken at 3-hr intervals
throughout the tidal cycle at stations near the outfall. These data will
help in locating the plume and in interpreting dye study results. More
detailed discussions of dye studies and fluorometers are included in
Feuerstein and Selleck (1963), Smart and Laidlaw (1977), Turner Associates
(1971), and Wilson (1968).
Navigation-Position Determination
Position determination is important for siting and returning to
sampling stations and for tracking drogues. Although several different
methods are available, the most commonly used methods for nearshore coastal
surveys are Loran C, electronic range-positioning systems with onshore
transponders, horizontal sextant readings from aboard ship, and theodolite
readings from shoreline stations.
Loran C is the government sponsored radio navigation system selected
for use in the coastal zone by the Department of Transportation. The Loran
C system has recently been implemented on the West Coast and in Alaska. The
previously existing chains on the East Coast, Gulf Coast, and the Hawaiian
Islands have been expanded and modernized to provide a complete network of
122
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Loran C staions for the U.S. coastal zone. Receivers automatically resolve
Loran C signals and display position with claimed accuracies as high as +_ 15
m (+_ 50 ft).
Portable electronic range-positioning systems are available from
several manufacturers and have claimed accuracies as high as +_ 1 to 3 m (+_ 3
to 10 ft). These systems are the preferred method for position
determinaton. Electronic range-positioning systems are limited to
line-of-sight measurements. The maximum range varies with the manufacturer.
However, most types have ranges sufficient for nearshore applications. It
should be noted that at very short ranges (<50 yd) electronic
range-positioning system errors may be disproportionately large.
Visual methods of position determination such as sextant or theodolite
readings are limited to use under conditions of adequate visibility. In the
sextant method, the position of the observer on-board a boat is fixed by
measuring two horizontal sextant angles between three charted objects, with
one object being common to both angular measurements. In the theodolite
method, the position of an object in the water (for example a drogue, or a
boat at a sampling station) is determined by simultaneously taking two
theodolite readings from separate shoreline stations. By knowing the
locations of the theodolite stations and the horizontal angles to the
observed object, the position is fixed. The accuracy of positions
determined with a sextant or theodolite varies with the precision of the
instruments, the experience of the operators, the horizontal distances
involved, the magnitudes of the angles, the strength of the three-point fix
(for sextant readings), the scale of the charts, and the accuracy with which
the positions are plotted.
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