United States
Environmental Protection
Environmental Research
Corvallis OR 97330
February 1980
Research and Development
Characteristics of
Nonpoint Source
Urban Runoff and
Its Effects on
Stream  Ecosystems


Research reports of the Office of Research and Development, U S  Environmental
Protection Agency, have been grouped into nine series  These nine broad cate-
gories were established to facilitate further development and application of en-
vironmental technology  Elimination  of traditional grouping was  consciously
planned to foster technology transfer and a maximum interface in related fields
The nine series are

      1   Environmental Health Effects Research
      2   Environmental Protection Technology
      3   Ecological Research
      4   Environmental Monitoring
      5   Socioeconomic Environmental Studies
      6   Scientific and Technical Assessment Reports (STAR)
      7   Interagency  Energy-Environment Research and Development
      8   "Special" Reports
      9   Miscellaneous Reports

This report has been assigned to the ECOLOGICAL RESEARCH series This series
describes research on the effects of pollution on humans, plant and animal spe-
cies, and materials Problems  are assessed for their long- and short-term influ-
ences Investigations include formation, transport, and pathway studies to deter-
mine the fate of pollutants and their effects  This work provides the techn cal basis
for setting standards to minimize undesirable changes in living organisms in the
aquatic, terrestrial, and atmospheric environments
This document is available to the public through the National Technical Informa-
tion Service, Springfield, Virginia 22161.

                                              February 1980
                  Donald B. Porcella
                  Darwin L. Sovensen
             Utah  Water  Research Laboratory
                 Utah State University
                   Logan, Utah  84322
               USEPA Order No.  CC81224-J
                    Project Officer

                   Kenneth W.  Malueg
                  Freshwater Division
      Corvallis Environmental  Research  Laboratory
                Corvallis, Oregon  97330
                CORVALLIS,  OREGON  97330
          „ •;  environmental Protection
          r-.c.?,ion V, Library
          230 South Dearborn Street
          Chicago, Illinois  60604


     This report has been reviewed by the Corvallis  Environmental  Research
Laboratory, U.S. Environmental  Protection Agency,  and approved for publica-
tion.  Approval does not signify that the contents necessarily reflect the
views and policies of the U.S.  Environmental  Protection Agency, nor does
mention of trade names or commercial  products constitute endorsement or
recommendation for use.
                      U4S. Environmental Protection Agency


Effective regulatory and enforcement actions by the Environmental  Protection
Agency would be virtually impossible without  sound scientific data on pollu-
tants and their impact on environmental  stability and human health.  Respon-
sibility for building this data base has been assigned to EPA's Office of
Research and Development and its 15 major field installations, one of which
is the Corvallis Environmental  Research Laboratory.

The primary mission of the Corvallis Laboratory is research on the effects of
environmental pollutants on terrestrial, freshwater and marine ecosystems;
the behavior, effects and control of pollutants in lakes and streams and the
development of predictive models on the movement of pollutants in  the biosphere,

This report surveys the literature on urban nonpoint source runoff in order to
formulate a basis for evaluating the impacts on benthic stream communities.
                                              Thomas A.  Murphy, Director
                                              Corvallis  Environmental  Research


     Literature on urban nonpoint source  runoff was  surveyed  to  determine
the magnitude of the effects  of that source  of contaminants to stream  eco-
systems.  Very little information was available on ecosystem  effects
although extensive literature on all  aspects associated  with  contaminant
loading was available.   However, urban NFS  runoff probably exerts  unique
effects on stream communities because of  its random  magnitude/impulse
loading of contaminants.  Because control of NFS runoff  is expensive,  it is
important to determine  its actual  impacts on stream  communities.  Ecological
literature provided a basis for evaluating  such impacts  based on benthic
invertebrate biomass and diversity,  measurement of community  primary produc-
tion and respiration, carbon  cycling, and variables  related to the  contami-
nant concentrations in  the stream.   We concluded that  a  stochastic  approach
for assessing the impacts would be most feasible for evaluating  impacts of
urban NPS runoff.

Foreword     .     .     .     .     .     .     .     .     .     .     .     .    iii
Abstract	iv
Tables	vi
Figures .............    vii
Acknowledgments   	     ix

   1.    Conclusions     ..........      1
   2.    Recommendations     .........      3
   3.    Introduction   ..........      4
             The Urban NPS/Stream System   	      4
             Urban NFS/Time and Space Factors   	      6
             Objectives     	      6
             Approach for Literature Survey     .....      6
             The Definition and Need for Study of the Effects of
               Urban Nonpoint Source Discharges to Streams     .     .      8
   4.    The Significance of Urban Runoff   .     .     .     .     .     .     12
             Sources and Concentrations  of Contaminants in Urban
               Runoff	12
             Stormwater and Pollutant Modeling  .     .     .     .     .     16
             Hydrology and Transport  .     .     .....     18
             Water Quality Variables  	     24
   5.    Ecological Concepts      ........     40
             Purpose of Ecological  Analysis of Urban NPS Runoff     .     40
             General Ecological Considerations  .....     40
             Categorization of Ecological  Variables  ....     42
             Selected Ecosystem Variables  for Stream Impact Analysis     45
             Runctional Groups and  Materials  Cycling ....     54
   6.    Case Histories of Random Event and Magnitude Impulse Impacts
          on Streams   ..........     58
             Flooding Effects on Fish Communities    ....     58
             Effects on Toxicant Spills     ......     58
             Stream Comparisons - Michigan ......     60
             Urban NPS Runoff Impacts on Stream Communities     .     .     63
   7.    The Analysis of Urban NPS Runoff Impacts on Stream Ecosystems     66
             A Probability Hypothesis for  Ecological Response
               Variables	66

References	71
Bibliography ............     87


Number                                                                  Page

   1     Range of Concentration in Urban  Stormwater [Wanielista 1978]      13

   2     Summary of Nonpoint Source Characteristics [Loehr 1974]     .     14

   3     Estimated Effects of Changes from Natural  Conditions  in
        Watershed use on Specific Water  Quality (WQ)  Variables .     .     25

   4     Concentrations of Contaminants in Urban Runoff    ...     27

   5     Loadings of Urban Runoff ........     28

   6     Freshwater Environmental  Criteria (Toxicity,  Irrigation,  and
        Others).  Lowest Value Cited Except for Variables Cited as
        Drinking Water Standards ........     29

   7     Precipitation Characteristics [from Loehr 1974]   ...     38

   8    Variables Identified with Properties that Describe the State
        or Condition of Ecosystems    .......     43

   9    Material Turnover Times in Terrestrial  Ecosystems [taken from
        Reiahle et al. 1975]	57

  10    Variation in Terrestrial Ecosystem Variables  [taken from
        Reiahle et al. 1975]     ........     57

  11     Comparison of Stream Quality Variables  for Different Rivers
        in Michigan [Ball et al. 1973]	61

  12    An Energy Budget for the Three Rivers [Ball et al. 1973]    .     62

  13    Range of Loading Rates Recorded for Storm Events in 1975    .     64


Number                                                                  Page

   1     The hierarchical  stream/watershed system that  produces  material
        flow to aquatic ecosystems     .......      5

   2     In the United States,  101  references  on  urban  runoff  impacts  on
        water quality were identified  .......      8

   3     Hypothetical  response  of stream community to random inputs  of
        toxic/stimulatory materials  as  exemplified by  urban NPS     .     10

   4     Intensity-duration-frequency curves  for  rainfall.   [Taken
        from Wanielista 1978;  used with permission of  Ann  Arbor Science
        Pub!.]	18

   5     An idealized  example of the  hydrograph and BOD and SS
        concentration (pollutograph) and load (loadograph) as a
        function of time.   [Adapted  from Wanielista 1978;  used  with
        permission of Ann Arbor Science Publ.]   .     .     .     .     .     19

   6     Variation of  flow and  water  quality  with time  during  a  runoff
        event.  (Taken from Randall  et  al. 1978; Denver Street
        Station]  ...........     21

   7     Effects of urban runoff on water quality of an hypothetical
        stream (runoff data interpolated from Figure 6 data)    .     .     22

   8     System classification  of transport of urban runoff to
        streams   ...........     23

   9     The relationship between streamflow  and  added  concentration of
        contaminant from loadings.   Levels of selected water  quality
        variable criteria added to show significance  (see  Table 6)   .     30

  10     Separation of processes and  concepts  for analysis  of  urban  NPS
        impact on stream ecosystem     .......     31

  11     Biological relationships of  the organism [from MaoMahon et  al.
        1978]	41

Number                                                                  Page

  12    The effects of environmental  fluctuations on steady state
        biomass in an ecosystem.   Three levels of persistent biomass
        are hypothesized based on nonvariant, variant,  and a highly
        variant environment wherein severe perturbations occur [from
        Reiohle et al.   1975]	46

  13    Comparison of the species diversities of the benthic macro-
        invertebrates between Stations 1  and 6, Green River, Green-
        field, Massachusetts \_DiGiano et al. 1976]   ....     65

  14    Hypothetical stream ecosystem effects of random impulse
        (urban NPS) loadings on ecological variables ....     67

  15    Percent survival decreases with more severe urban runoff
        events    ...........     69

  16    There is experimental error for system variables which adds
        to the variance caused by response to stochastic inputs     .     69


        The following persons  contributed  their knowledge  and  skills  to  this
report:  Nancy W.  Winters  (ecologist),  Edward F.  Cheslak  (ecologist),  Scott
Reger (ecologist), Garry Laughlin (biologist), David  S. Bowles  (hydrologist),
and L. Douglas James (hydrologist).   Garry Laughlin  helped extensively in
preparation of the literature  review.   Edward Cheslak provided  excellent
review of the ecological chapters and David Bowles extensively  revised and
refined our thinking in Section VII.   The  rough draft typing was  done  by
Sherri Barber, Norma Schiffman, and  Mardyne Matthews.   The final  copy  was
prepared by Kathleen E. Bayn.   Our special  thanks to  these people and  to
Kenneth W.  Malueg, project officer,  whose  initiative,  sympathy, and patience
helped us put this report  together.

                                SECTION  I


     Although the literature contains extensive evidence  of widespread urban
runoff problems in the United States, there  are few reports that describe
the effects of urban runoff on variables related to stream biota and their
interrelationships.   Consequently,  it is difficult to assess the actual
impacts of urban runoff on stream ecosystems.   Also, the  literature is inade-
quate to determine the magnitude of the urban  nonpoint source (NPS) runoff
problem.  Such an inventory would require an assessment of river basins  in
the U. S. and their associated pollutant sources.   However, using the number
of annual literature citations and  specific  studies reporting the relative
magnitude of pollution sources an an indicator, it is apparent that urban
NPS runoff is one of the most serious problems affecting  water quality.

     Because of the unique characteristic of contaminant  input to streams
from urban runoff, the effects on stream biota could be extensive.   Contami-
nant inputs are severe and have significant  impacts on water quality varia-
bles.  However, the temporal and spatial patterns  of contaminant concentra-
tion in the stream could result in  differences in  community response when
compared to steady inputs such as occur with point source waste discharges.

     It is necessary to determine the impacts  of urban NPS runoff on streams
in order to set priorities to ameliorate the problem.  The widespread
occurrence of water quality degradation by urban runoff,  the high cost of
treating or instituting best management practices  (BMP) and the unclear
relationship between water quality  variables and community variables are
factors which must be considered.  Control of urban NPS runoff is directed
at the input phases of the hierarchical urban  runoff system:  watershed,
hydrology, and transport.  Treatment and BMP would affect the input of
contaminants that controls water quality.   Ecological variables are affected
by water quality.  Therefore, the ecological effect of contaminants derived
from urban NPS runoff must be determined before deciding  on the proper
priorities and methods for institution of expensive NPS controls.

     The following specific conclusions that relate to these statements  were
derived from the literature:

1.   Stream water quality can vary  markedly  both temporally and spatially as
     a function of stream flow and  quality and watershed-hydrologic-transport

     a.   Watershed factors include percent  impervious area, traffic and
          residential densities, industrial  activity, street sweeping activity.


     b.   Hydrologic variables include seasonal patterns of storm intensity
          and duration, annual precipitation, interval between storms.  Also,
          the receiving stream depends on the same hydrologic variables.

     c.   Transport factors include the collection and discharge system,
          presence of detention basins, separate or combined  sewers.   Trans-
          port processes and the hydrologic regime of the  stream combine to
          determine the stream concentration for a given load of a contami-

2.    Several  stream water quality variables are affected significantly by
     urban runoff:  BOD5, refractory organics,  suspended sediments,  salinity,
     Pb, Zn.   Although BOD5 is a commonly measured quality variable,  it is
     not very representative of the total organics load (COD:BOD ratios can
     approach 10:1) produced by urban runoff.  Other water quality variables
     (nutrients, bacteria, Cu, Cd, ...) are significant in specific cases.

3.    The relationship of stream biota response  to the highly  variant stream
     quality is  unknown.  However, effects on biota have been observed in
     comparative studies.  Biomass, diversity,  and P/R ratios are affected.
     Generally benthic invertebrate biomass and diversity decreased.   In
     some cases  P/R ratios approached 1.0, indicating greater autotrophic
     productivity; in other cases where organic loads were decreased, P/R
     was less than control streams.

4.    Response of biota is hypothesized to be a  function of stream concentra-
     tions of contaminants.  Based on an assessment of the literature and
     evaluation of selected case studies, a set of ecological variables was
     defined that would permit comparative evaluation of the  ecological
     effects of urban NPS runoff.  Comparison with an absolute standard is
     not feasible at this time and evaluation of effects must be based on
     upstream/downstream, before/after, natural/affected stream comparisons.

     The set of ecological variables would include, but not be restricted to,
the following:  biomass and diversity of invertebrates, gross primary
production (autochthonous and allochthonous) and system respiration, carbon
dynamics, and dissolved oxygen dynamics.  Habitat description and selected
water quality variables would also be determined.  Hydrologic variables must
be determined.  Methods suggested by whipple et al. [1978] should be applied.
Stochastic models such as Padgett et al. [1977] and Slatkin [1978] should be
used to provide sampling times, locations, and  intervals.

                                SECTION  II


     Experimental  studies to elucidate the effects  of urban  NFS  runoff
impacts on stream  communities should  include  two  steps:

1.   A synoptic survey of streams  throughout  several  regions of  the  U.S.   The
     streams should be regionally  significant,  have comparable reaches or
     tributaries that are relatively  natural  and  impacted  by urban NPS
     runoff, and adequate historical  stream data  and hydro!ogic  data should
     be available  or obtained.   Analyses  of comparable stations  during
     critical  (defined by dynamics of natural  community) seasons of  the year
     should include benthic invertebrate  biomass  and diversity,  productivity
     and respiration measurements, carbon turnover  times,  and characteriza-
     tion of water quality variables.   Experimental  error  of measurements
     should be determined.

2.   An assessment of community response  based  on the stochastic formulation
     described in  Section VII should  be performed and compared to effects  of
     similar contaminant loads  produced by steady (point sources) rather than
     impulse loadings.  Some related  hypotheses that should  be tested include
     the following:

     a.   Impact on communities is a  function  of  water concentrations of
          contaminants.   Such an hypothesis assumes body burdens are a
          function of water concentration which may not be true  in a highly
          variant  system.

     b.   Carbon turnover time  can be assessed  simply and  results related  to
          energy flow as well as material  cycle.

     These hypotheses are the overall  assessment  of community response and
stochastic systems should be evaluated in controlled or laboratory streams
to minimize the error and confounding associated  with natural stream systems.

                                SECTION III


     Stream ecosystem impacts  of materials  in  urban  runoff relate  to  several
specific and unique features that distinguish  those  impacts  from other linked
systems.  The purpose of this  study is  to determine  if these unique  features
result in ecological  differences that can be distinguished by appropriate
ecosystem variables.


     By definition, materials  in urban  runoff  originating from nonpoint
sources (NFS) and combined sewers are excluded from  direct consideration.
These NFS materials are transported by  water generated as a  stochastic
(random, but characterized by  time of the year)  hydro!ogic event,  are affected
significantly by temporal  and  spatial factors  that have preceded the  hydro-
logic event, and impact aquatic ecosystems  as  a  discrete, largely  receiving-
system-independent input.   Urban runoff impacts  have similarities  with other
aquatic ecosystem impacts  including the relationships between activities and
materials produced, the materials themselves,  and interactions within the
aquatic ecosystem that are not a function of the input.  The system watershed
can be visualized as consisting of separate but  somewhat interrelated phases
or processes:  (1) the production of waterborne  materials in the watershed
which can enter the aquatic ecosystem,  (2)  the hydrologic event, (3)  the
transport system, and (4)  the  aqueous environment and its community  (Figure  1).

     The watershed phase of the system  is considered to be the natural system
plus a set of activities that  produce varying  quantities and types of mate-
rials, depending on the types  of activities.  Typical concentrations  and
mass loadings of materials indicate that there are regional  differences in
urban runoff.  The differences can be ascribed to geography  and varying
activities taking place in the watershed.  These activities  include  industry
type (air pollutants), street  sweeping  policy, urban density, and  traffic.

     The hydrologic phase  consists of the water-producing storm or precipita-
tion event.  The factors that  affect runoff quantity are the extent  to which
runoff from paved areas is collected by a storm drainage network,  and ante-
cedent moisture conditions as  determined by recent storm activity  and specific
variations in area and activities.  These factors cause variations in flow
and pollutant concentration.  Concentrations of pollutants vary with  type  and
intensity of the different activities that can and do occur  in the watershed.
However, the hydrologic event  and runoff intensity and duration control the
mass of pollutants picked  up in the runoff, their concentration, and

1. Watershed Phase
 Watershed Activities
 Land Uses
  Random hydrologic
  event that produces
  adequate runoff to
  reach aquatic receiving
  system.  The variables
  include temporal and
  spatial factors such as
  time since last storm,
  intensity of storm in
  time and space
Effects on
   Material load =

   FLOW *
         4.  Ecosystem
       Fate in
 2.  Hydrologic  Phase
                                   3.  Transport
                               In stream
Figure 1.  The hierarchical  stream/watershed system that produces material
           flow to aquatic ecosystems.
ultimately the mass flow into the stream.   To some extent, depending on
relative runoff and stream flows, dilution of materials in the receiving
stream depends on the magnitude of the hydrologic event.

     The transport system is that phase of the hydrologic event wherein the
runoff is transferred to the stream.   It is important to  distinguish between
typical situations such as pipe inflows to streams, retention basins, and
recharge basins, because aquatic community response varies considerably with
the type of spatial and temporal mechanisms in which materials enter streams.

     The stream ecosystem presents many challenges for analysis:  stream
morphometry, geological differences that result in variations in sedimenta-
tion and water quality, flow (hydrological) variations, other cultural
impacts (waste discharges and variable land uses and their associated
pollutants), and stream community variables.   The community variables change
with environmental factors such as flow, water chemistry, stream morphometry,
and fish and wildlife management.  These phases of the urban NPS material

input to streams can be visualized as a matrix of specific variables that can
be associated with stream community response and pollutant inputs.


     As a generalization, pollutants from urban NPS are qualitatively the
same as pollutants from any other source and impact aquatic ecosystems as
the specific contaminant independent of source.  However, in a quantitative
sense, the intensity and frequency of input varies considerably and randomly
on a seasonal basis.  Such inputs are described as stochastic and their
impact on stream communities is hypothesized to result in a unique  community
which has the stability, resistance, and resilience to survive such impacts.
It is important to define variables that reflect those impacts, the range of
intensity of those impacts, and frequency of occurrence.


     The purpose of the project was to perform a literature review  to deter-
mine the state-of-the-art of assessing the ecological  impacts of urban
nonpoint source runoff on streams.  The review would result in a report on a
national assessment of urban NPS runoff impacts on biota and ecological
system variables.  Although there are extensive results on water quality
impacts, there is little information that relates to impacts on biota and
stream ecosystem variables.  Consequently, the literature search was expanded
to permit the development of concepts relating to impact assessment that
could be evaluated by specific research projects.

     The following objectives were established:

1.   Provide an overview of current literature on watershed variables, and
     hydrologic and transport processes of urban NPS runoff.

2.   Review the recent literature on effects of urban  NPS runoff on water
     quality in streams.

3.   Discuss concepts of ecosystems function and how contaminants affect
     variables related to ecosystem function.

4.   Discuss actual case histories of urban NPS impacts on stream ecosystems.


     The intent of the literature review was to obtain information  on
biological/ecological indices of freshwater stream ecosystem impacts of
urban runoff.  Consequently, several fields of study were identified as
relevant:  stream ecology, biological/ecological indices, urban runoff and
its quality and quantity, and spatial and temporal characteristics.  Because
of the breadth of the review, some reduction in citations was necessary.  For
example, a series of annual reviews from 1973 to 1978 of studies of urban
runoff and combined sewers contained a total of 907 citations and most of
these describe the chemical and hydrological properties of transported
contaminants and their sources [Field and a series of co-authors 1973, 1974,

1975, 1976, 1977, 1978].   Other studies and reviews indicated a similar
situation.   Therefore, we concluded that an exhaustive review of the entire
literature would be inappropriate and repetitive.   The references cited
herein were selected to illustrate example processes and properties of urban
runoff and their effects  on stream ecosystems.   Similarly, ecological litera-
ture covers a broad array of subjects, not all  of which are specifically of
value, and specific references were selected to illustrate concepts.  We are
indebted particularly to  an unpublished review by Hendvix [1979] and textbooks
by Wan-ielista [1978] and  by Linsley and Franzi-ni [1964] for background

     The survey of literature was initiated using computerized bibliographic
data bases which are commercially accessible (Lockheed Information Services
or Systems Development Corporation).  In addition, a search on the appro-
priate subject areas (urban runoff, aquatic ecosystem response to pollutional
stress) was performed on  the Water Resources Scientific Information Center
(WRISC) bibliographic data base.  Commercially available data bases searched
were COMPENDEX (The Engineering Index - monthly/annual, 1970-current), NTIS
(National Technical Information Service, 1964-current), BIOSIS PREVIEWS
(Biological Abstracts., Biovesearoh Index, 1972-current).  SCISEARCH was used
exclusively to search for papers citing previously published papers with
major emphasis on the topics of interest.

     In addition to computerized systems, the contents of journals published
since 1969 which were throught most likely to publish pertinent papers were
searched manually.  These included Water Research, Journal of Water Pollution
Control Federation, ASCE  Index 1970-1974, 1975; Proceedings of the Environ-
mental Engineering Division ASCE; Proceedings of the Hydraulics Division,
ASCE; and Proceedings of the Irrigation and Drainage Division, ASCE.  The
contents of the last three proceedings were searched from 1976 to current
since previous publications would have been included in the ASCE Index.
Ecological journals searched manually included Limnology and Oceanography,
Ecology, Ecological Monographs, Arkiv fur Hydrobiology, and Journal Fish.
Res. Bd. Canada.

     We used the following key words with appropriate key word strings for
the computer searches:  ecosystem, community, freshwater, reservoir, stream,
river, estuary, lake, response, recovery, stress, impact, bioindicators of
pollution, land use and urbanization, water quality modeling, urban runoff.

     The computer searches provided few useful  references.  They included
979 citations of which 133 (14 percent) were applicable to the urban runoff
system.  Of these, only 2 dealt with ecological impacts while 54 were reviews
or dealt with models and  hydrology/hydraulics.   Of the remaining 77 articles,
46 were oriented to urban runoff and 31 dealt with standard water quality
analyses.  This pattern was similar for reviews of urban runoff.  Therefore,
we concluded from the results of the computer searches, literature searches,
and discussions with workers in the field, that essentially all urban runoff
research is oriented toward assessment of watershed, hydrologic, hydraulic
and water quality variables, and functional relationships.  Except for the
few reports discussed in  the section on case histories, essentially no
research on ecosystem impacts has been conducted.   This is despite the unique

mode of impact that urban NPS runoff has on streams.

     However, there is significant interest in urban  runoff subjects.   We
were able to identify 119 reports that relate to urban runoff, water quality
or ecological impacts of urban runoff.  Although most U. S. studies were
concentrated in the Great Lakes and Middle Atlantic States, reports from all
regions of the continental U. S. plus Canada, European and other countries
were reviewed (Figure 2).  This broad interest indicates the widespread need
for assessing ecological impacts of urban runoff.  Perhaps ecological  impacts
were neglected because a testable hypothesis for analyzing urban runoff
stream impacts had not been formulated.   This report  represents a contribution
toward the development of such a hypothesis.


     Urban environments are not typical  of the histor-'cil < ^environmental
events that occur during the evolution of natural commuiit is that reside in
specific ecosystems.   Processes occur in urban systems tha  permit only the
most tolerant and resilient species to grow and reproduce.  Streams receiving
inputs from precipitation-generated runoff that transport a wide variety of
materials that have either no effect or are toxic or  stimulatory, will
develop communities that are unique to the system.  Within the context of
Figure 2.  In the United States, 101 references on urban runoff impacts on
           water quality were identified.


describing this problem,  we define the following limits  to the urban/stream
hydrologic system:   (1)  Urban environments are systems that are covered by
large amounts of impervious surfaces of roads and buildings, and are incorpo-
rated and have a population greater than 2500 [U.S.  Department of Commeroe
1975].  Thus, small, as  well  as large, population centers are included, i.e.,
cities and towns.   (2)  Streams are limited to freshwater rivers, creeks,
etc. and are small  and/or large, varying from first order to the higher stream
orders (6-12).  (3)  The aqueous transport of materials  deals with all  non-
water matter as contaminants, but not necessarily pollutants.  A pollutant
is any factor that, in the sense of Maokenthun [1975], affects the "integrity"
of the aquatic environment.  Urban contaminants are usually the same as
those from wastes and natural sources, but there are specific differences in
their impact resulting from the temporal and spatial mode of input.

     Urban nonpoint sources are many and varied and their inputs to aquatic
systems are characterized by the stochastic nature (random but characterized
according to time of the year) of hydrologic events, the carryover effects
of previous events, and  specific activities in the urban watershed.   Conse-
quently, receiving water communities are impacted by a wide variation in
concentration (input load and dilution capacity vary) of a broad spectrum of
physical factors, chemical toxicants, and biostimulatory compounds.   Control
of such impacts includes the three major approaches of applying best manage-
ment practices (BMP), requiring nonsteady state criteria for materials, and
biological or ecological  criteria for ecosystem response to urban inputs.
The last approach deals  with the integrity of the aquatic ecosystem; however,
the aquatic ecosystem response is least understood and requires considerable
research; it has a great effect on the development of BMP and criteria.

     Individual organisms and the community of a specific ecosystem will
likely respond to urban  runoff as the concentration of material (temperature,
suspended solids, toxicants, nutrients) rises from a relatively  steady  level
to a peak value and falls off, roughly analogous to a flood peak hydrograph.
The occurrence of these  hydrologic events is random.  The peak value depends
on total loading, timing and duration of the loading input and the dilution
capacity of the receiving water.

     Urban systems have  significantly less lag time for hydrologic events
than do natural or rural  systems because of the relatively large amounts of
impervious surfaces.  Also, peak flows and concentrations could be greater
and the rising and falling rates faster than in natural  streams.  The
response of biota will depend greatly on uptake rates, growth rates, motility
(relates to recolonization, attraction and avoidance), and rates of repair.
Interpreting the ecological effects of urban nonpoint sources will depend to
a large extent on biotic measures which integrate the long term impacts of
these stochastic events.   A conceptual approach to visualizing the probable
effects of such events shows that the wide variety of probable outcomes can
be narrowed if better data and understanding of the system processes and
components are obtained  (Figure 3).  Models exist for evaluation of hydrologic
inputs in this conceptualized system [for example, Bovee and Milhous 1978].

     In 1974 the International Hydrological Decade published a report of a
study by its subgroup on the effects of urbanization on  the hydrological

^     4-
o     ^
  -I CO
  < LJ
  o or



   (controlled  by LOAD/FLOW)
              Long Term Steady
              State Based on
Figure 3.   Hypothetical response of  stream community to random  inputs of
           toxic/stimulatory materials as exemplified by urban  NPS.
environment  [international HydTologioal Decade 1974].   Case studies from
the Federal  Republic of Germany,  the  Netherlands, Sweden,  USA, and the USSR
were presented along with special  topic studies of illustrative  importance.
A summary of findings indicated that  the self purification capacity of the
lower Rhine  River had been reduced by 30 percent due to toxicants.  Studies
in Sweden showed that, compared to treated wastewater, storm water runoff had
more SS,  less BOD, and less phosphorus and nitrogen.  In addition, flows from
separate  sewers had as great a pollution potential as  combined sewer overflows.
It was also  concluded that, due to the sudden and brief nature of urban
runoff and the large number of outfalls, the reduction in  pollution from urban
runoff will  be very costly, possibly  more costly than  wastewater treatment.

     A questionnaire to all national  committees for the IHD identified six
priority  topics for research on the effects of urbanization on the hydrologic
cycle. They were:

1.   Changes in surface runoff caused by urbanization, amount and quality
     (including sediment), for storm  water sewerage and urban streams;

2.   Research on quantity and quality of runoff in terms of precipitation
     occurrences in experimental  catchment areas;

3.   Soil moisture and groundwater, quantity, and quality relationships and


4.   Water demand forecasting, including technical,  economical, social, and
     political  aspects, and quality of source considerations;

5.   Water quality effects related to groundwater arising from direct dispo-
     sal of wastes and to indirect pollution from disposal of solid waste

6.   Effects of wastewater and sludge on the natural purification capacity of
     receiving waters and their aquatic life, taking into account the
     processes involved.  [International Eydrologiaal Deoade 1974].

The first five topics relate to the watershed, hydro!ogic and transport
phases of the urban NPS runoff system.  Topic 6 is related to the subject of
this report and is probably the least studied but most significant at this

     Our literature review showed that ecological effects of urban NPS runoff
have largely been ignored and we believe this has occurred because controls
are usually oriented toward contaminant production.   However, because of the
high cost of controls for urban NPS runoff, it is essential  that relationships
between urban runoff and ecological impacts be determined.  When these
relationships are defined, proper priorities for establishing control measures
can be instituted.

     This report is directed at NPS, although frequently the impact of urban
watersheds on stream ecosystems is the result of other categories of pollu-
tant sources.  An example of such a major source is  combined sewer overflow.
Combined sewer overflow behaves very similarly to any urban runoff event but
will not be discussed directly herein because this review and analysis is
confined to NPS.  However, we realize that combined  sewer effects may occur
in specific instances and may not have been separated in the reports
reviewed.  Also, ecosystem impacts can be viewed as  independent of the source;
thus, ecosystem and community concepts relate to combined sewer overflow as
well as direct urban runoff.  The concepts developed herein apply to any
pollutant input that has random, relatively discrete, single event properties.

                                SECTION  IV



     Water pollutants from urban  watersheds  are  produced  by  a  variety  of
activities (sources) that are affected by  location  variables and activity
intensity.  Attempts to  identify  regional  characteristics  have been  largely
unsuccessful.   For example, Bradford [1977]  tried to  normalize his data to
perform this analysis.   A statistical  summary  of data available through 1972
relating land use to materials on urban streets  available  for  removal  by
storm water, did not show significant  patterns or relationships.  The  author
felt that the data were  inadequate to  show  relationships  because of  different
methods of collection and analysis.  These  differences introduce variance
which is not representative of the actual  processes taking place.

     The types of pollutants are  somewhat  typical of  many  point sources
(Table 1).  However, unlike many  point sources such as wastewater effluents,
concentrations in NPS waters are  extremely  variable (Table 2).  One  way of
looking at the effects of urban runoff is  to compare  the  effects of  urbaniza-
tion.  For example, MoGriff [1972] concluded that urbanization increased the
volume of runoff and the size of  the flood  peak, and  decreased the lag time.
Effects on water quality were threefold:   (1)  sediment load  increased,
(2) groundwater recharge was minimized; hence, long term  stream flow augmen-
tation and dilution was  reduced,  and (3) eutrophication resulted with
consequent effects on DO and related variables.

     Urban watersheds typically consist of transportation  surfaces,  loading
dock surfaces, building  surfaces, and  vegetated  surfaces.  These surfaces
vary considerably in their permeability.   Roads  and sidewalks, loading
surfaces, and buildings  are essentially impervious  and all of  the precipita-
tion becomes runoff with minimal  lag time.

     The Soil  Conservation Service identifies  four  types  of  soils that affect
the amount of runoff {Kent 1973].  Type A  has  the  lowest  runoff and  is
composed of sand and gravel, type B has moderately  low runoff  and is mostly
sandy, type C has moderately high runoff typically  being  composed of a shallow
soil layer containing some clay,  and type  D has  high  runoff  and is essentially
impermeable because of the high clay content.  Bare soil  at  constructions
sites is a major problem in urban areas, particularly as  a sediment  source.

     In most urban situations, bare soil  is  negligible and  permeable  land
surface is generally a vegetated  surface such as gardens,  parks, roadsides,



Variable                          Low                     High

BODs                               1                   700 mg/£
TOC                                1                   150 mg/£
COD                                5                 3,100 mg/£
SS                                 2                11,300 mg/£
Total Solids                     200                14,600 mg/£
Volatile Total Solids             12                 1,600 mg/£
Settleable Solids                  0.5               5,400 mg/£
Organic N                          0.01                 16 mg/£
TKN                                0.01                  4.5 mg/£
NH3N                               0.1                   2.5 mg/£
N03N                               0.01                  1.5 mg/£
Soluble P04                        0.1                  10 mg/£
Total P04                          0.1                 125 mg/£
Chlorides                          2                25,000 mg/£*
Oils                               0                   110 mg/£
Phenols                            0                     0.2 mg/£
Lead                               0                     1.9 mg/£
Total Coliforms                  200               150 x 106/100 m£
Fecal Coliforms                   55               110 x 106/100 m£
Fecal Streptococci               200                 1.2 x 106/100 m£

t Table adapted and used with permission of publisher, Ann Arbor Science Publ
* With highway deicing.


.- -• §3
oo v ooo , 7T
o» — i 0) rs O O
OO O f-l-^O CM ^ O»
odd d o" o" d """^
O 0
i/> *° °,
. , ^ ^ (N . "^
1 1 V O If
"2 * SS" - |J
• • 1/1 o* -*i
odd d -^
o o
0 O
~> ° 2"d
71 I - 7 i
~ ' ^- ' ~. oo
2 S2 oo
-" " f^ O
o —
in O
2 ~
7 IS 7 oo
^ in o o
T3 2i ai ^ to "O
S=C g § M^
§1 _2 . 2^ |2
•Ilell le^ll ll«i
and vacant lots with ground cover.   Thus,  the runoff from vegetated surfaces
can approach the quantity of precipitation only if the ground is saturated by
previous precipitation or if the soil  is of an impermeable type.  Soils
covered with vegetation are quite permeable and surface runoff usually does
not occur unless extensive periods  of precipitation saturate the soil.
Interflow (subsurface flow) and delayed runoff are important effects of such
vegetated surfaces.   The relative mixture of impermeable and permeable
surfaces depends on  cultural, socio/economic, geographic, geologic, climatic
and urban planning/management variables.

     Deposition of materials on urban surfaces is highly dependent on other
related activities.   For example, vehicles directly deposit  (1) tire
particles, (2) oil,  grease, and fuels, (3) spilled materials from transported
goods, and (4) fallout from vehicle exhaust emissions.  The amount of these
materials depends on the kinds of vehicles (automobiles, light trucking,
heavy trucking) and  the intensity of use by each category of vehicle.  Other
materials that can be found on transportation surfaces include wastes from
pets and birds, solid wastes (trash collection loss, litter, garbage, leaves,
and other vegetation), and industrial and power plant air pollutant fallout.
Street sweeping and  washing reduce the amounts but may be relatively ineffec-
tive in removing some kinds of pollutants {.Sutherland and MoCuen 1978].  The
percentage of land surface covered by roads and sidewalks controls the rela-
tive impact of these activities on urban runoff.

     Barkdoll et al. [1977] concluded that dustfall effects on urban runoff
quality were very contaminant specific.  In a Knoxville, Tennessee watershed,
mass flow of COD, Hg, Cl, As, and POit-P in stormwater was closely related to
dustfall levels for  these contaminants.  They also concluded that minerals
and solids have relatively constant concentrations from storm to storm while
heavy metals, nutrients, and COD have decreasing concentration with increasing
runoff.  Another significant conclusion was that pollutant removal is
primarily a function of total runoff and only slightly affected by runoff
intensity (depth of runoff).

     Sartor and Boyd [1972] found that the total amount of accumulated
contaminant material showed a relationship to elapsed time since the last
rain or street sweeping.  There was wide variance in the amount of contami-
nant material observed (290-3,500 Ib/curb mile), but it averaged about 14,000
Ib/curb mile in the  cities tested.   The largest portion of polluting sub-
stances was associated with fine solids in the street surface contaminants.
Rainfall intensity,  street surface characteristics, and contaminant particle
size were the controlling variables for the rate of rainfall wash-off but
street sweeping was  primarily aesthetic because it did little to remove
contaminants that would be carried in runoff, i.e., fine grained particles.
Catch basins were found to remove coarse grained inorganic materials, but
were not effective in removing fine particles.

     Loading dock areas and buildings can serve as significant sources of
pollutants depending on the type of construction material and activities
that occur.   Spilled chemicals, oils and grease, and other materials at
loading docks occasionally provide massive inputs to streams.  The density
and type of loading  activity will control  the probability of such events.


Similarly, buildings and the activities associated with buildings are distri-
buted at changing density within an urban area.   A comparison of runoff
quality from different types of buildings indicates that contaminant load is
generally greater in urban than in commercial areas, which is greater than in
single family residential areas \_Randall et al.  1978].   This observation may
be more closely associated with traffic than building type and density,
although it is difficult to say with complete surely.  Traffic volume and
contaminant load is greater on highways than city streets, which in turn are
greater than for rural road systems [Wanielista 1978].


     The concepts presented in this section, although elementary and somewhat
incomplete, are intended only as the introduction to a complex subject.  The
concepts discussed are those that relate to assessing the impacts of urban
runoff on receiving streams.  There are detailed models in the literature.
Many are evaluated by Wanielista [1978], but only two, the U. S. Army Corps
of Engineers model (STORM) and the USEPA model (SWMM), are applied by him as
illustrated examples.  Most models use flow weighted average values and are
of some value for management but do not relate to the stream impacts of urban
runoff as described herein [Wu and Ahler>t 1978].  The example publications
cited in the following paragraphs illustrate this problem; however, the
utility of specific models such as STORM and SWMM for generating management
questions and possibly some ecological questions cannot be excluded.

     A simulation model for stormwater management (SWMM) was developed by
Lager1 et al. [1971].  The model simulates both stormwater quantity and
quality and is capable of representing combined wastewater overflow phenomena.
The model was validated for the Baker Street combined sewer basin in San

     Hoyden et al. [1972] developed a method using computer modeling techni-
ques of predicting runoff management problems and their effects on a small
urbanizing watershed.

     Using a modification of EPA's SWMM, Roesnev et al. [1972] evaluated the
effects of land use changes on the quantity and quality of urban runoff in
the City of San Francisco.  Under the simulation of changing a 305 acre park
to multiple residential housing, large increases were predicted for watershed
outflow, suspended solids mass emission rate and concentration, and BOD mass
emission rate.

     Graham et al. [1974] developed equations useful in predictive modeling
to estimate imperviousness and specific curb length for urban watersheds.
The equations use Census data, including households per acre, population per
acre, and employment per acre.

     Shih et al. [1975] used a simulation approach to model urban runoff
hydrology for small  (<20 mi2) urban watersheds.  The modeling method
developed estimates of runoff rates and associated confidence limits.  Data
inputs are obtained from topographical maps  (watershed area, slope, channel
slope and length), soil survey maps (soil type and approximate minimum


infiltration capacity),  and National  Weather Service publications (rainfall
intensity-duration frequency).

     McElroy et al.  [1976] compiled a handbook of loading functions of
pollutants from nonpoint sources which included urban runoff.  Wu and Ahlert
[1978] recommended homogeneous-land-useand statistical  synthetic approaches
as accurate and practical  methods for storm runoff load prediction.

     Phamuon and Fok [1977] developed a digital computer model for urban
runoff simulation.  The model  was checked using data for a Hawaiian small
urban watershed.  Model  performance was judged satisfactory.  They were
criticized by Gold-ing [1978] particularly for the overall approach and
instrumentation for measuring flow.

     Rovey and Woolhiser [1977] developed a computer model for estimating
runoff from urban watersheds which relies on a series of planes and channels
(kinematic cascade) to simulate a complex watershed.  The model was success-
ful in predicting runoff from a watershed near Northglenn, Colorado.

     Diskin et al. [1978]  developed a two-component parallel cascade model
for predicting urban runoff.  The model input is the total rainfall hyetograph
for the watershed.  The two elements of the model calculate runoff from the
impervious and pervious area of the watershed in parallel; then the summation
of the two elements represents the final output.

     Jewell et al. [1978]  developed a method for calibrating quantity-quality
coupled urban stormwater management models.  The method calibrates the
quantity model first and then the quality model.  Calibration is done prima-
rily by adjusting watershed characteristic parameters for the quantity model,
and by adjusting pollutant buildup rates for the quality model.  Average
conditions measured for several storms are recommended as calibration data.

     Work by Whipple et al. [1977] indicated that it is not valid to assume
that pollution from urban runoff increases with elapsed time since a previous
rainfall.  Earlier assumptions were erroneously made because there had been
insufficient total pollution loading data available for individual storm
events.  The authors criticize other assumptions (e.g. , the effect of street
sweeping) in recent modeling efforts (i.e., stormwater management model of
EPA) and call for a wider questionning of the relationship between storm
characteristics and pollution loadings.  In answer to the criticism of
Whipple et al. [1977], Field [1978] pointed out that the USEPA's stormwater
management model (SWMM) includes nonlinear storm runoff-washoff functions
which were not taken into account by Whipple et al.  [1977].  He also points
out that the SWMM dry-weather pollutant accumulation function is based on
real data and is not arbitrary.  The SWMM model is not a "complete solution
for urban runoff problem assessment... but is an accepted tool in developing
solutions to this complex area of environmental management."

     Sutherland and MoCuen [1978] used a computerized urban nonpoint pollu-
tion management model (developed in 1976) that estimates the accumulation,
removal by rainfall, and removal by street sweeping of eight pollutants
(total solids, volatile solids, BOD5, COD, TKN, N03, PO,,, total heavy metals)


to simulate the effects of street sweeping  operations  in the Washington, D.C.
area.  They concluded that street sweepers  did  not ordinarily remove more
than 33 percent of the pollutants accumulating  on  streets because of (1) the
inability of sweepers to remove small  particles (which contain most of the
pollutants) and (2) the less than optimal street sweeping operation procedures.


     Storms at a specific location and season tend to  occur randomly with
respect to time, space, and intensity  (Figure 4).   Probabilities of occurrence
for precipitation events that have a specific intensity (depth of precipita-
tion per hour) for a specific duration are  called  intensity-duration-frequency
curves [Wanielista 1978].  Different localities will have different patterns.
             10 0

                              SANTA FE, NEW MEXICO

              1 I
              e t
              o c
             0 4
               5        II
                                  30 41
                               MIAMI,  FLORIDA
3   4 5
                                                       i » It 12  It 24
               5        10
                                                       i > 10 12  II 24
Figure 4.  Intensity-duration-frequency  curves  for rainfall.   [Taken from
           Wanielista  1978;  used with  permission  of Ann Arbor Science Publ.],

The precipitation events  themselves are probabilistic events;  that  is,  they
are random, mutually  independent events.   One can predict a long  term
average with a certain  probability of success but the occurrence  of specific
events is not predictable and  follows a binomial distribution.  Runoff  is
affected by these time, space,  and intensity variables, plus other  variables
associated with the watershed.   These considerations have led  to  concepts
that are specifically related  to management [Wanielista 1978]:  the hydro-
graph, the pollutograph,  and the loadograph (Figure 5).  It should  be
mentioned that this terminology is not universally accepted, although the
terms illustrate useful concepts [tfh-ipple et al. 1977].  Chemograph has been
used in place of pollutograph.

     The hydrograph is  a  relationship between flow (§) and time (t}; flow
increases to a peak value and  then decreases exponentially depending on storm
   5.10   2700

   4.54   2400

   3.96  I 2100


   3.40  § 1800

           2.83 §  1500

 - 2.27 g 1200

I      t
   1.70 *~ 900

                                                             36.3 —
18.1  co
                                                 54 .5
                                                           45 o:

                          40      80     120    160   200   240
Figure 5.  An idealized example  of  the  hydrograph and BOD and SS concentra-
           tion (pollutograph) and  load (loadograph) as a function of  time.
           [Adapted from Wanielista 1978;  used with permission of Ann  Arbor
           Science Pub!.].

and watershed characteristics.   The pollutograph is the concentration of the
specific contaminant (CC^) in the runoff;  the loadograph is the mass flow
(Q*CCi = SL}.  An example of a pollutograph/hydrograph shows how TOC, SS, and
nutrients vary with time and flow (Figure  6)  and illustrates several

1.    The "first flush effect" where concentrations  are high>r in the initial
     runoff because of buildup of materials between storm events and subse-
     quent dilution effects of storm water.

2.    The transport of particles requires sufficient velocity to pick up
     particles and keep them in suspension.  This is especially apparent in
     comparing the N and P compounds where TKN and  total P are associated
     with particulate and the other nutrient forms  are ior, ized.

3.    While water quality variables are measured in  the runoff, it is the
     impact of those variables on the aquatic community that requires
     analysis to interpret impacts on stream ecosystems.

     Although many investigators have looked at the storm water hydrologic/
quality system, the impacts of the storm water on receiving waters have not
been carefully evaluated.  Using the data  in Figure 6 [Randall et al. 1978]
for suspended solids in storm water, the suspended  solids concentration
in a receiving stream can be calculated (assuming behavior as a conservative
substance) to illustrate the impacts of a  storm event on that stream (Figure
7).  The average stormwater flow over the  90 minute calculation interval is
0.117 cms, while the peak flow was 0.150 cms and the minimum flow was 0 cms.
The storm runoff characteristics are described by the hydrograph, the loado-
graph, and pollutograph.  However, these events are greatly affected by the
relative amount of dilution in the stream itself.

     The maximum impact of the storm runoff occurs  with a stream flow of
1 cms (approximately 8.5 times the flow of the runoff event) and at approxi-
mately 50 minutes after the beginning of the runoff event.  Increasing the
dilution factor by a factor of 5 (approximately 40 times the runoff flow)
reduces the impact by a corresponding factor.  Further dilution eventually
produces a condition where essentially no measurable impact occurs (stream-
flow > 100 cms).  It is evident that a stochastic approach is needed to
assess the impacts of the urban runoff on  water quality in the stream.  The
stream flow varies extensively with time and is greatly affected by storm
events.  The combination of stream variation with storm runoff variation and
pollutant concentration variation requires that the stream impact be assessed
using a stochastic approach.  The components of the aquatic community are
impacted by those variables as events which may have a short term or a long
term impact on their growth or survival.

     Although the stormwater runoff, as shown in Figure 6, is measured as if
it were entering the stream in Figure 7 at a specific and discrete point
over the time interval and it is assumed that there is instantaneous and
complete mixing in the stream, in reality urban runoff is transported by a
wide variety of hydrologic systems.  These systems can be categorized as
surface or subsurface, point input or lateral input, delayed inflow or direct


                                           SUSPENDED SOLIDS
                                                            .... JOB
                                                              JO) I
                   5 _
Figure 6.   Variation of  flow and water  quality with  time during  a  runoff
            event.  [Taken  from Randall  et al. 1978;  Denver Street  Station.]

                                                     Qs=I.O cms

 Additional  60
 Solids, mg/l


                                                                       o   m
                                                                       U-   co

                                                                       0.10 400
              0    10   20   30   40   50  60

                                 Time , minutes
Figure 7.  Effects  of urban runoff on  water quality of an  hypothetical
           stream  (runoff data interpolated from Figure  6  data).

inflow (Figure 8).  Although many examples occur, the four shown are probably
the most typical kinds of inputs to streams receiving urban runoff.  Example
No. 1 shows a storm drain entering a retention basin.  Systems of such basins
appear to be a valuable means of managing urban runoff [refer to Haan and
DeVore 1978].

     For example, Dale [1978] discussed the efforts of the Metropolitan
Sanitary District of Greater Chicago (MSDGC) to control flooding and pollu-
tion from urban runoff through the construction of a system of tunnels and
reservoirs to hold runoff water until it can be treated.   The system,
currently under construction, will serve a 971 km2 (375 mi2) area in the
center of metropolitan Chicago.  When completed, the project is projected to
save $750 million/year in flood and water pollution damage.  The system will
help protect Lake Michigan and allow local waterways to meet "fishable and
swimmable standards."
                  Types of urban runoff transport systems
                             Surface     I
                             Subsruface  II
                             Point       A
                             Lateral      B
                             Delayed Inflow
                             Direct Inflow
      Storm Drain

Figure 8.  System classification of transport of urban runoff to streams.


     The delayed inflow of example 1  could represent a diversion of storm
drain water into a treatment plant as might occur for first flush influents.
Cherkauer [1977] reported on the ability of a series of lakes in a small
stream system in the suburbs of Milwaukee, Wisconsin, to modify the intensity
and quality of urban runoff.  The storage capabilities of these lakes have
a positive impact by lowering the peak discharge and "flashiness" of surface
runoff and increasing the length of time a stream is above baseflow levels
during a storm event.  The lakes also even out concentrations of dissolved
solids including Na and Cl which are  deposited during road salting.  Another
example of a delayed flow mechanism is the use of wetlands for treatment of
urban runoff waters \_Hiokoak et at.  1977].  The wetlands were effective in
upgrading the quality of urban runoff if application was properly managed,
and apparently no impacts to the wildlife or vegetation (type or abundance)
occurred as a result of the project.

     Thus, the discharge of urban runoff into streams can take a wide variety
of inputs, further complicating the idealized picture shown in Figure 7.
Impact on aquatic communities is dependent on the temporal aspects as
exemplified in Figure 7 and the spatial aspects as exemplified in Figure 8.
All of these aspects combine with the watershed characteristics to produce
the input to the aquatic ecosystem.  The aquatic ecosystem is impacted by
those variables associated with the urban runoff as it is diluted in the
streams.  These variables include flow, various water quality parameters, the
character of the stream in relation to the urban runoff and the composition
of the aquatic community itself.

     For purposes of simplification,  we define short term and limited impacts
as those areas within the stream where no permanent and/or significant effect
on the aquatic community occurs.  There are certain types of impacts that
might be limited in space but have a  marked effect on the aquatic community.
An example would be the development of a barrier that prevents movement of
higher organisms past a specific point due to water quality problems.  DO
deficits can function in this manner to prevent fish migration.  Another
example is the loss of an essential component of an aquatic community, which
may affect survival of other components.  With these constraints in mind, we
can ignore certain minor impacts that result from water quality changes
that are limited in area and time.  In many respects a "mixing zone" concept
can be considered as the zone where these small impacts take place \_USEPA


     As stated in the Introduction, we describe the flow input phase of the
urban runoff process in terms of the watershed and the hydrologic and
transport systems.  The material input phase of that process involves
specific physical, chemical, or biological variables that represent contami-
nants picked up in the runoff and transported into the stream.  A listing of
selected variables important in analyzing urban runoff and the activities
that produce them indicates that considerable variation in activity control
measures would be required  to improve water quality across the range of
variables selected (Table 3).  Although many studies have been performed on
these variables, they are not always measured equally well nor at similar






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ti w a 

     Concentrations (Table 4) and mass loadings (Table 5) of many of these
variables indicate that runoff concentrations can be higher than raw sewage
and have significant impact on a receiving stream.   The values in Tables 4
and 5 were obtained by arithmetically averaging data summarized in the cited
papers.  Even with relatively low loadings and high dilution of highly toxic
materials (mercury and some of the chlorinated hydrocarbons), environmental
and other criteria [abstracted from USEPA 1976] can be surpassed for a period
of time (Table 6).  Assuming an urban area of 10 km2 and mass input within a
single 24-hour period, the average concentration in a stream would follow the
relationships shown in Figure 9.

     The interaction between watershed activities,  water quality variables,
and stream ecosystem responses is best understood by separating the different
components of the system.  An abstract example (Figure 10) shows that each
phase of the urban NFS runoff process affecting the stream ecosystem can be
separated in terms of its impact on the stream response variables.

     Thus flow and flooding and specific water quality variables such as
salinity, suspended solids, organics, bacteria, and nutrients interact with
specific aquatic physiological and ecological functions.  These functions
include respiration and photosynthesis and the concepts of structure, such
as food webs, succession, and trophic levels.  In this section we discuss
selected literature examples of chemical, physical, and biological variables
traditionally associated with water quality [see Pennaek 1971, Gakstatter et
al, 1977, Ott 1978a, b, for general review and discussion of water quality
and various indices].

Flow and Flooding

     Urbanization results in inert !sed imperviousness of the watershed which
in turn enhances the intensity and volume of runoff from the watershed \Ceeh
and Assaf 1976, Gundladh 1976, Hossain et al. 1978].  These factors often
link urbanization with increased flooding of streams draining these watersheds
[Doehring and Smith 1978].

     Under conditions of uncontrolled development in the Menomonee River
watershed in Wisconsin, Walesh and Videkovioh [1978] found increases of up to
4.5 times in predicted 100 year flood stages.

     The hydrologic alterations of urban development in coastal areas were
studied by Cech and Assaf [1976].  In the Houston,  Texas, area, the quantity
of urban runoff was 3-5 times more than non-urban runoff, and 2-2.5 times
higher in the smaller Beaumont area.


     Most severe  salt impacts result from road deicing.  Timing is critical
due to season, freeze-thaw cycles, storm events \_Soott 1976].  Jodie [1975]
reported on the quality of storm water from urban freeways in Milwaukee,
Wisconsin, and showed that salt concentrations were generally high with surges






TDS Tot. P

1100 7.3
20 < 0.02
190 1.10
2 3

Cl" N03-N Alkalinity

14 270
< 1 20
< 3 93
2 2





Mn Ni

0.842 0.066
0.132 0
0.408 0.018
7 7
Pb Zn Fecal
Col i form
100 m£
< 2
0.598 0.694 240000
01080 0.035 9
0.271 0.211 36500
77 12
1 Brwnlee et al.  [1970].
2 Whipple et al.  [1974a].
3 Veils et al.  [1975].
11 Randall et al.  [1978].

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 Figure  9.   The relationship between streamflow and  added  concentration of
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Effect of
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                  Process  controlling variables
                  Net flow of  materials;  varies in magnitude with
                  time and space; varies  in  significance according
                  to magnitude and type of material,  uses of
                  stream, and stream  variables.
Figure  10.  Separation of processes and concepts  for analysis of  urban NPS
           impact on stream ecosystem.


during the winter and spring.   Although salt input from deicing may enter
streams indirectly, it can usually be almost totally accounted for.

     Hawk-ins [1976] reported that some of the salt from street deicing in a
New York urban watershed was stored in the groundwater.  This would delay
input to streams and also continue input after seasonal road salting was
completed.  Approximately 82 percent of the applied salt could be accounted
for in the watershed drainage.

     Judd [1970] showed how significant salts could be when studying the
effects of runoff on shallow,  glacially-derived,  First Sister Lake in Michi-
gan.   Studies were made from the winter of 1965 through the winter of 1967.
There were heavy snows in both  1965 and 1967 and, therefore, road salts were
used  extensively.  Due to input of road salts from the surrounding area, the
lake  stratified and failed to turn over during the winter and spring thaws of
1965  and 1967.  The lake did turn over during the winter of 1966.  The effects
of the salt for the three years were:

                     Sampled in April, ymho/cm at 18°C

                     1965           1966          1967

Surface (0 m deep)    442            339           361
Bottom (6 m deep)     690            366           648

Also, complete depletion of dissolved oxygen in the bottom layer occurred in
1965  and 1967.


     One of the most significant problems associated with urban NFS runoff
is suspended solids (SS), inorganic and organic materials that cause turbidity
and sedimentation problems.  In Tucson, SS in runoff decreased as residential
areas became more developed \M-Lsoh-L and Dhasmadhikari 1971].  Frequently,
"first flush" concentrations of SS are higher than raw sewage and later
stages are equivalent to secondary effluents [Corberg 1977].  When combined
with  the flow produced in runoff, the impact of runoff pollutants is greater
than  the secondary effluent for the same area in Sydney, Australia \_Corberg
1977].  One runoff event in Sydney lasting less than an hour that resulted
from  13 mm of rain over a 131  km2 area produced 1150 kg SS, 100 kg BOD5, 13 kg
PO^,  and 11 kg NH3.

     Ellis [1976] studied the water quality of urban runoff in the Silk
Stream catchment which includes 3323 km2 of North London, England.  Rainfall-
runoff lag times were only 20 to 30 minutes.  The first flush carried
between 350-3000 mg/& SS, with the peak of solids and flow approximately
coinciding.  Volatile solids followed a similar pattern.  Inorganic sediments
make  up 45-70 percent of the SS.  The bulk of the sediments ranged between
0.1-0.5 mm diameter and showed a log normal distribution with a tendency
toward bimodality as the distance of transport increased.  The stream had a
DO sag where sediments accumulated.  Those sediments were high in Cd and Pb.


     In presenting some of the findings of the International  Reference Group
on Great Lakes Pollution from Land Use Activities,  Judd and Carlson [1978]
point out that urban areas contribute high loadings of phosphorus, sediment,
heavy metals, toxic organics, and chloride.   This pollution is from storm and
combined sewers as well as runoff.

     The application of statistical procedures by Edith [1976] in selecting
land use and water quality in New York river basins indicated that river
concentrations of suspended solids, nitrogen, but not phosphorus were related
to land use.  High density residential land use was an important factor
affecting suspended solids in rivers.  However, the relative contribution to
the river of point and nonpoint sources in suspended solids pollution from
urban areas could not be estimated in the study.

     Guy and Jones [1972] argue that to prevent excessive sediment loads to
streams caused by the ongoing process of urbanization, erosion control is an
important aspect of design.  Proper controls can minimize this input.
However, the origins of SS are not always due to erosion but may be related
to accumulated particulates that vary with housing density and other factors.

     Whipple et al.  [1978] provided data illustrating the complexity of
studying urban NPS runoff effects:
                                      Suspended Solids from a Modern
                                       Multiple Family Housing Area

              Date                              kg/storm

             7/23/76                               20.1
             8/06/76                              299.
             9/10/76                              189.
            10/26/76                                1.82

To obtain these data, the flow contaminant variable had to be sampled
throughout the storm event; single samples did not suffice.  Also, the land-
scaping and impervious surface provided excellent erosion control so the SS
resulted from accumulated particles.  The authors demonstrated that, in
general, modern multiple family units provided more contaminants per unit
than did single family housing.


     Although Pb is the principal metal associated with urban runoff because
of its use in automotive fuels, Zn and other metals are frequently found at
deleterious levels.  The association of Pb and Zn in sediment cores with the
onset of urbanization signals the potential impacts of urban runoff on
aquatic ecosystems [Chris tens en et al. 1978].

     Bryan [1974] reported 0.33 kg/km2 day of lead discharged from the 106.7
acre (432 ha) urban drainage basin in Durham, N.C.   In one instance, dissolved
lead concentrations were thought to be great enough to interfere with BOD
measurements.  However, subsequent observations revealed that the lead was
associated with suspended solids and had no apparent effect on BOD.


     Oliver &t al.  [1974] in studies of urban Ottawa, reported that amounts
of chlorides in snow and runoff from snow melt were related to street deicing
practices.  However, lead came from leaded gasoline and was found to be
tightly bound to fine particulate matter in the snow.  Thus, if plowed snow
were dumped away from waterways, the vast majority of lead from this source
would be retained at the dump site with the particulate matter.

     Newton et al.  [1974] reported that lead deposited on city streets from
automobile exhaust resulted  in significant concentrations of lead in runoff.
In the Oklahoma City area, the authors predicted an average lead concentration
in runoff of 0.23 mg/£.   An  adjusted average concentration of lead in snow
and ice collected from a heavily traveled street in Oklahoma City was 3.2 mg/H.

     Wilber and Hunter [1975] reported on heavy metal contributions from
urban runoff, and presented  data from two drainage areas in Lodi, New Jersey.
Seven individual storm hydrographs were analyzed and first flush effects were
found to be highly significant.  Pb, Zn, and Cu were the major heavy metals,
accounting for 90-98 percent of all the metals observed (Pb, Zn, Cu, Ni, Cr).
Pb:Zn ratios suggested the sources of these metals in runoff and precipita-
tion may be similar.  With regard to the relative contributions of metals in
stormwater runoff, direct precipitation, and secondary treatment plant
effluent, runoff accounted for as much as 86 percent of the total.

     Also, Wilber and Hunter [1975] studied urban runoff of heavy metals
discharged into the Saddle River near Lodi, New Jersey.  Base flow metal
concentrations were quite variable, probably due to industrial use variations
in the sub-basins.  Confirming their previous results, stormwater runoff
showed a "first-flush" effect for heavy metals and the distribution in preci-
pitation was similar to that in runoff, i.e. , Pb, Zn, were dominant.  Bottom
scouring of unconsolidated sediment was found to be important in metals'
transport by rivers.

     The chemical relationships of heavy metals in streams are being deve-
loped [Vuoeta and Morgan 1978].  The effects of temperature, pH, and other
ions on the concentrations of metallic ions in water explain the equilibrium
concentrations available for biotic uptake or other reactions in stream
ecosystems.  Davis and Leokie [1978] illustrate how these metals could be
accumulated in sediments.

Organic Pollution and Dissolved Oxygen

     A major question about the significance of NPS pollution emerged with the
application of stream quality models for analyzing the effects of point
sources on stream dissolved oxygen (DO) resources.  Earlier editions of these
models showed that relatively good prediction of observed data was obtained
by applying the models under well-defined stream conditions.  However,
results were poor when NPS pollution became important; this indicated that
NPS pollutant impacts were significant to the DO resource.  Many investiga-
tions then showed the impacts of NPS organics as measured by Biochemical
Oxygen Demand (BOD) on the stream DO [Corder 1977, Misahi and Dhasmadhikari
1971, Whipple et al. 1974a,  1974b, 1976a].


     The models, based on the Streeter-Phelps [1925] equations, have been
improved and applied to practical situations by a variety of investigators
so that the methodology is essentially a textbook procedure \_Nemerow 1974,
Thomann 1971].  There are specific processes peculiar to urban runoff, e.g.,
the sudden impulse of load that requires analysis [Smith and Filers 1978].  A
recent review of the water quality modeling technology illustrates some of
the refinements in the approach [_Grenney et al.  1976].  Future refinements
include the aspects of randomness and experimental error by applying DO
models that involve stochastic techniques \Mdlone et al.  1979].

     Application of modeling and statistical approaches for analyzing urban
NFS runoff has added insights to the material input phase of the process.
Rickert et al. [1975] used a model of DO in the Willamette River, Oregon, to
evaluate the effects of reducing carbonaceous and nitrogenous oxygen demand.
They concluded that nonpoint sources contributed 46 percent of the total
oxygen demand during the dry weather season of 1974.  Thus, point source
treatment alone would be an incomplete method for eliminating severe oxygen
demand problems in the Willamette.

     Often, however, the question of organic inputs to streams has not been
separated from the problem of combined sewer overflow as is frequently
observed in the eastern United States.  Gates [1975] concluded that combined
sewer overflows were important point sources of BODs, and that significant
amounts of BODs were contributed by nonpoint sources in the study area.  Also,
Pennine and Perkins [1978] reported that current overflows contributed a
minimum of 0.66 million kg of BOD and 1.97 million kg of suspended solids to
the James River (Virginia) each year.  Lindholm [1976] studied pollution from
combined sewers and used mathematical models to analyze the problems of BOD5
discharge, retention basin design, setting of storm overflow, and the size
of the secondary clarifier in the wastewater treatment plant.

     Organic loads include easily degraded materials (measured by BOD) and
carbonaceous materials that are not as readily degraded (COD and TOC).  For
example, Thompson et al.  [1974] showed that urban runoff in Lubbock, Texas,
contained COD/BOD ratios of 9.5:1.  Typical raw sewage ratios are 2-4:1
\_Eokenfelder and Ford 1970].  This reflects the synthetic nature of much
of the urban NPS runoff as contrasted to fecal materials in domestic waste-
waters.  Mackenzie and Hunter [1979] showed that aromatic compounds in
Delaware River sediment were derived from slow-degrading crank-case oil, a
component of urban runoff.  Wakeham [1977] showed similar sources for Lake
Washington (Washington).

     However, most NPS organic material loading has been assessed as BOD.  In
non-urban areas (including single family housing), Chippie et al. [1974a]
found that during rainy periods, organic (BOD) loading from unrecorded
sources increased up to 10 times more than during a typical dry period.  They
also found that cropped areas contributed as much BOD loading in the New
Jersey area they studied as did single family housing.  Industrial and non-
industrial urban areas contributed significant amounts of BOD.  The weighted
mean annual BOD loading for an industrial area was about 235 kg/km2/dy with
storm runoff being the controlling influence.  An urban residential area
contributed an average of 64 kg/km2/dy.


     Yu et al.  [1975] investigated organic pollution for seven small  New
Jersey watersheds, three in urban areas.  They found substantial organic
pollution (BOD5) from an urban-industrial and an urban-residential  area with
light industry.  This BOD load could be greater than or equivalent  to the
load from secondary sewage effluent.  Loading from storm periods was  more
than 10 times the dry weather load.  Agricultural  areas could contribute as
much or more organic load as urban-residential areas.  Frequency distribution
of BOD concentrations were plotted and were projected to be very useful in
modeling efforts.

     Rimer et al. [1978] applied land use planning concepts to stormwater
runoff in the Piedmont Region of North Carolina (Triangle J Council of
Governments, 208 Study).  Generally, NPS pollution increased with increasing
impervious area.  Business districts did not follow the pattern because of
minimal land disruption and the practice of street sweeping.  Authors found
that DO was an important water quality variable for stream management and
that the other variables (BOD5, COD, SS, TKN, N03-N, TP, TOC, heavy metals
at selected sites) were less important.  However,  these variables were the
input form of the toxicant that affected the stream DO.

     Whipple et al. [1976] studied stream pollution from urban watersheds in
New Jersey.  They concluded that for residential areas BOD loads from
unrecorded sources would range from 9-13 g/person/day.  In heavily
industrialized/commercialized areas, the loads would be greater.  The
unrecorded pollution sources must be considered if pollution control  plants
are to be cost effective.

     Sullivan [1974] of the American Public Works  Association denies Whipple's
assertion that the APWA report underestimated BOD due to possible runoff
prior to street sweeping or catch basin cleaning.   He also points out that
the information cited by Whipple et al. [1974a] is specific to the  Chicago

     Although the sources of organics that contribute to DO depletion in
streams are varied (pet and vegetable litter, fuels, oil and grease,  combined
sewers, garden sources), several unique sources that could on occasion have
great and sudden impact on streams have been noted.  Thompson et al.  [1974]
observed that discharges from fire fighting runoff were highly polluted with
COD concentrations as high as 1740 rng/A and suspended solids concentrations
of 670 mg/£.  Sahultz and Comerton [1974] studied the Montreal  International
Airport and estimated that projected increases in air traffic would increase
the organic load in storm sewers from that airport by the year 1985-86, to
8100 kg BOD/day.  The use of aircraft deicers (diluted ethylene or propylene
glycol) provided most of this load.

     Weibel et al. [1966] showed that organo-chlorine pesticides could
attain significant levels in urban runoff from a residential-light commercial
section of Cincinnati, Ohio.


     Indicators of bacterial pathogens have been used to assess the potential


danger to human health of urban NPS runoff.   Olivieri et al.  [1977] found
high levels of recoverable pathogens and indicator organisms  in urban runoff
from Baltimore; Pseudomonas aeruginosa (103-105/100 m£), Staphylocooous
aureus (10°-103/100 mi); Salmonella and enteroviruses were found at levels
of 10° to 10V10 liters of urban runoff.  Shigella, although  not recovered
due to culture process interferences, may have been present.   The authors
recommended that because of relatively low densities of pathogens, and
current regulations prohibiting use of water polluted with urban runoff, that
it would not be cost effective to disinfect the large quantities of urban
runoff.  They recommended, however, that combined sewer overflows (10 times
higher in pathogens) be disinfected to reduce the load for subsequent
culinary water treatment plants and contact recreation use hazard.

     Geldve-ioh et al.  [1968] compared the bacteriological character of storm-
water runoff from city streets, a suburban business district, a storm sewer,
and a wooded hillside adjacent to a city park in Cincinnati,  Ohio.  Highest
seasonal mean values of fecal coliforms were measured in street gutters
(47,000/100 m£) and runoff from a business district (40,000/100 m£) in the
autumn.  They concluded from fecal coliform/fecal streptococci ratio (FC/FS)
data that fecal contamination was mostly from cats, dogs, and rodents
(FC/FS-0.7) rather than human (FC/FS-4.4) sources.  Fecal indicators and
pathogens survived storage in stormwater longer at winter temperatures (10°C)
than at summer temperatures (20°C).

     Davis et al. [1977] observed a "first flush" effect in all hydrographs
from rural and urban drainages in the Houston area.  Fecal coliform and
fecal streptococcus concentrations were higher in the more developed areas.


     Nutrient problems in aquatic ecosystems are primarily directed toward
lakes and impoundments and the problems associated with increased plant
productivity as a result of eutrophication.   Lake restoration by point source
nutrient diversion is ineffective when NPS inputs are not controlled [Emery
et al. 1973].  The effects of nutrients in streams have not been well
documented; however, undoubtedly increased productivity would result from
added nutrients \_Ball et al. 1973].  Such nutrients can be supplied
extensively by urban NPS runoff.

     Loehr [1974] summarized the characteristics of nonpoint sources of
pollution from various land uses; agriculture, forest lands,  and urban runoff
(Table 7).  The relative contributions of urban runoff with municipal/
industrial contributions to aquatic ecosystems for nutrients  were compared;
a range of 0-87 percent for nitrogen and 0-89 percent for phosphorus was
obtained in wastewaters while urban runoff contributed 0-48 percent and
0-57 percent, respectively.  Thus, urban runoff contributed significant
quantities of nutrients in certain locales.

     Weibel [1969] discussed the sources and types of urban drainage in the
light of eutrophication potential.  He pointed out that rainfall from the
Cincinnati, Ohio area contained enough inorganic nitrogen (0.69 rngA) and
total phosphorus (0.08 mg/£) to support algal blooms.  Urban runoff adds



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significant amounts of N and P to the rainwater, and combined sewer overflows
may be very high in nutrients (3.0 mg/£ total P).  Various methods of preven-
tion and treatment are discussed.  Most include storage with subsequent use
of the water resource, or routing to secondary treatment plants.

     Wenster,  et al.  [1975] showed that success for the eutrophication control
program in Swedish lakes by point source phosphorus removal depended signifi-
cantly on the relative contribution of urban NPS runoff nutrients.

                                 SECTION V

                           ECOLOGICAL  CONCEPTS


     The analysis of the effects of urban NPS runoff on  stream  ecosystems  is
a special case of the general  problem of the ecological  analysis  of changes
in environment, communities, and specific organisms.   Patrick [1949] and
many others have argued that biotic and  ecologic changes are  the  most
accurate for assessing physical  and chemical  changes on  ecosystem health
because of their integration and direct  response to the  complex melange of
chemical and physical variables.  In this analysis we attempt to  develop
concepts that are holistic rather than directed at a specific variable such
as DO or a critical organism.   It is a macroscopic rather than  microscopic
view of ecological function and structure, thus the defined variable
boundaries tend to be inexact and at times lump more than one process,
trophic level, species, etc.  Although definition of holistic properties  of
the macroscopic system that we define to be  an aquatic ecosystem  is a
relatively new field of endeavor, it should  be possible  to define a set of
variables useful for interpreting ecological  changes in  streams.   We intend
to provide a selected list of variables  against which "before/after" or
"upstream/downstream" effects can be observed.  The defined target variables
should aid in explaining and perhaps predicting community response to random,
impulse-type, variable magnitude environmental changes in streams.


     "An ecosystem is a set of organisms and inanimate entities connected by
exchanges in matter or energy" \MaoUdhon et al. 1978].  Biological relation-
ships of organisms and larger organizational systems can be shown using a
hierarchical approach (Figure 11).  Perturbations of all types  impact
ecosystems at the level of growth and reproduction of individual  organisms
[Wooduell 1970].  Interactions in communities or the biosphere consist of
either matter/energy impacts on organisms or organism impacts upon organisms
viewed in terms of effects on growth and reproduction \_MaoMahon et al. 1978].
Growth and reproduction respond to concentration of energy or material and
can be viewed as an uptake process or an enzymatic process [Reynolds et al.
1975, Chen and Selleok 1968, Fitzgerald 1969, Sprague 1971, Eppley and
Thomas 1969, Miohaelis and Menten 1913, Monod  1949].  This can be represented
as follows using the symbols of Monod [1949] for growth controlling materials
 S...S  ):
        n                           s2,

               KINGDOM     COMMUNITY
           DIVISION or PHYLUM


               ORDER       POPULATION

         2     FAMILY

              SPECIES         DEME




    of the ORGANISM

I  Physiologicol-Anatomicol

2 Phylogenetic

3 Coevolutionory

4 Matter-Energy Exchange
                    SUBCELLULAR STRUCTURE

                       SUBATOMIC PARTICLE

Figure 11.   Biological relationships of the  organism  [from MaoMahon et al.

     Because communities are made up of a set  of populations (all the
organisms of a  taxon) that in turn are composed  of  smaller assemblages (deme)
and the individuals themselves (organisms),  the  growth-promoting effects
(removal  mechanisms would be treated similarly)  of  materials and energy
inputs on the community can be viewed as the resultant  (yt) of the specific
effects on  growth generalized to the level  of  a  population (e.g., Patrick
                         y = g(vi,  U2, y3,...,  y)
The population  is the minimal size capable of analysis.  Community variables
relating to  effects on growth of populations  are  still  largely hypothetical
relationships.   Diversity estimates, biomass, turnover, keystone species,

relate to these hypotheses.  Probably the most successful  such demonstrated
relationships for aquatic ecosystems is the summer chlorophyll a - springtime
total phosphorus relationship in lakes as defined by Vollenweider in 1968
and many others since then.

     Chemical and physical laws apply to ecological  systems and are useful
for evaluating structure and function.  For example, in terms of the system
role of energy processing, the laws of thermodynamics can be applied with
reasonable accuracy at our present level of understanding of ecosystems and
of irreversible thermodynamics.  Although ecosystems act  primarily as energy
processing systems, rates of processing are controlled by essential elements
and by temperature.  Consequently, ecosystems could  be very inefficient
energy processors in order to maximize rate controlling nutrient concentra-
tions by rapidly cycling nutrients [Reiohle et al.  1975].

     Energy processing and nutrient cycling can be considered fundamental
attributes of ecosystems and community function and  structure would be
optimized to maximize these functions.  As an ecological  system, there
should be a set of variables that allows an essentially complete description
of the ecological state of the system.  Several key  macroscopic variables
have been identified as exemplifying important properties of ecosystem
structure and function (Table 8).  Although the listed variables are quite
general, they can be easily applied to specific situations with selected
specific measurements.  If these variables (Vy) are  visualized in terms of
growth, they can be related to growth regulating processes as follows:
Unfortunately, very few of these variables have been applied in urban stream


     There are several ways and many levels of categorizing the vast number
of biological, ecological, and environmental  processes and interactions and
the variables that measure them.  The physical or abiotic environment
contains those chemical and physical factors that are affected by or affect
organisms (biota).  The community is a set of populations of organisms which
function in the abiotic environment.  There are spatial  and temporal factors
involved in the interactions between biotic components and the abiotic
environment.  Although both concepts are confined to a specific and defined
space, function in a community is measured over a period of time while
structure is measured at a specific time (instantaneous).  There is some
overlap between these concepts.  For example, energy processing is a
functional process that takes place in a time scale between structurally
defined components.  The ecosystem of interest in this report is the fresh
water stream ecosystem with its fresh water biota and the abiotic components
of the stream and of the stream substrate and watershed.  The fresh water
biota can be characterized as composed of taxonomic groups or community
groups.  Examples of taxonomic groups include the various classes of
organisms (taxa).  Community groups can be classed on the basis of habitat
(such as, benthic, planktbnic organisms), or function within the system (such


                      Variables* (measured per standard unit area)
   I.   Environmental





                  Volume, pressure, abiotic mass flow rates, system
                    attributes (homeostasis-feedback, storage, loss)
Diversity, number, mass, chlorophyll a


C, N, P, ..., total mass

Change in diversity, number, weight, competition,
  resilience, resistance

Energy transfer, efficiency, P/R ratios

Elemental turnover times, P/R ratios, limiting
*  Can be applied at any or all levels:  community, trophic, population,
   organism, dominance, keystone, guild, (functional groups) or habitat.

as groups that have similar energy/matter processing characteristics; for
example, producers, consumers, decomposers; or autotrophs and heterotrophs
\Hendvix 1979].

     MaoMahon et at.  [1979] imply that habitat categorization is not a useful
concept for evaluating ecosystem structure or function.  In reality habitat
is a controller; it defines the community limits and types of organisms but
does not relate uniquely to community structure and function.   For example,
benthos refers to the organisms that inhabit the bottom of an aquatic
ecosystem.   The structure and function of the benthos is not dependent on
where they reside but on their niche.  Perhaps changes in habitat would
result in different possible niches but community structural and functional
relationships would be related to those niches not the habitat.   Consequently,
it is incorrect to sample a specific habitat to estimate community structure
and function.   The structural/functional  component of interest should be
analyzed and if the component occurs in a specific habitat, the  habitat
sampling would be valid.   Benthic macroinvertebrates would be a  valid
sampling if the processes studied are totally contained therein.

     Stalnaker [1979] and others in his group have developed a scenario for
describing change in spawning habitat with changes in flow and stream
morphometry.  This scenario is seasonally affected by hydrology  and life

cycle.  Given enough time, the resultant spawning frequency and fish produc-
tion that result from change in habitat produce population distributions,
community structure and function that are unique to that habitat and that
set of niches that result from the changed habitat.  If these arguments are
valid, response variables should not be habitat related.  For the remainder
of this report only niche variables (community structure and function)  will
be considered as valid response variables.

     Impacts of materials that enter fresh water ecosystems are of two  types.
There are direct impacts where direct linkages occur between the factor
being studied and a target biota.   Indirect impacts are those where the
biotic response is due to secondary interference with some target variable.
Examples of direct impacts would be direct toxicity to fish or stimulation
of algal growth by nutrients.   Indirect impact would include the effect on
zooplankton populations resulting from removal of zooplankton predators and
inhibition of primary productivity due to light transmission interference
from suspended sediments.

     Contaminants in freshwater ecosystems affect structure and function
because of their differential  effects on the species that compose the biota.
Sensitive species might increase or decrease depending on type, concentration,
and phenological relationships between species cycles and contaminant input.
Relationships between species  might be altered secondarily, with further
repercussions among the community depending on the significance of the
contaminant event and the affected species.

     Organisms would respond to NPS urban runoff contaminants produced
according to the scheme outlined previously in Figure 1; a hierarchy of
variables could then be defined to allow analysis of their impacts in the
stream.  Ideally one would assess urban runoff impacts using an ecosystem
model that would provide as output those critical response variables
identified as controlling some perception of ecological quality.  Although
examples of ecosystem models have existed for some time [NSF biome models,
e.g., Park 1978, Isvaelsen et al.  1975, Orlob 1975, also see selected articles
in Middlebrooks et al. 1973],  they are probably not suitable for analyzing
specific stream ecosystems because they were constructed with other objectives
in mind.  However, these models with suitable modifications may be useful
for analyzing the general case of urban runoff in streams.

     Models of streamflow, habitat, water quality variables, and many other
chemical and physical attributes of aquatic systems have been used to evaluate
the impacts of changes in stream inputs, thruputs, and outputs but so far
these models have not been linked to biotic attributes.  Quality variables
have been related to standards, beneficial uses and organism requirements,
but the direct linkage of abiotic and biotic variables has not been
accomplished.  Usually semi-quantitative comparisons of controlling
variables or of "before/after impacts" on biotic response variables have been
the limit of efforts in understanding stream ecosystems.


     The management of stream quality depends on the analysis of the effects
of water quality changes on stream communities.   Historically, this analysis
has been oriented toward perception of stream conditions and corresponding
biological {.e.g., Thomas et al.  1973], chemical, and physical indicators.
The most noted example of using indicators is the Saprobien system of
Kolkuitz and Mars son [1908, 1909, Sladaaek 1973, for recent review].  The
approach of linking problem perception to chemical/physical variables has
achieved the greatest acceptability in relating  productivity in lakes to the
nutrients, N and P \_Sawyer 1947, Vollenweider 1968, Dillon and Rigler 1975].
Generalized cause/effect relationships have not  been developed between biota
and chemical/physical variables in streams.

     The hypothesis that concentrations of contaminants affect ecosystem
response variables is testable and should result in functional relationships.
The problems in testing that hypothesis include:  (1) selecting appropriate
response variables, and (2) defining an appropriate variable that relates to
the concentration of the contaminant (discussed  in Section VII).  Ecosystems
process matter and energy and the impact of urban NPS runoff contaminants
(inputs of materials and chemical energy) on communities in stream environ-
ments is largely a question of the flux of matter and energy in the ecosystem
(Table 8).  Bartsch and Ingram [1966] reviewed and criticized attempts by
pollution control agencies to assess communities in stream reaches by
indicator organisms and other such environmental indicies.  Similarly, the
use of chemical/physical variables has been severely criticized.  More
recently, holistic ecosystem concepts have seemed to offer more value in
evaluating biotic and community response to stream perturbations such as
urban runoff.  These include biomass, diversity, and energy and material
turnover rates.  In this section we select a set of response variables based
on previous, successfully applied variables plus a minimally redundant set
from Table 8 to represent matter and energy flow.  In this way we expect to
minimize the economic dilemma of accounting for  every particle of the
ecosystem while avoiding omissions caused by the broad brush of generalized


     The biomass of a stream community is not as apparent as that of terres-
trial communities; in fact, except for fishing we generally prefer stream
communities to be relatively inconspicuous.  However, for every ecosystem,
Reichle et al. [1975] define a potential maximum biomass based on average,
steady state, peak or other parameter that is primarily related to energy,
elemental (nutrient) supply and temperature (Figure 12).  Perturbation or
random fluctuations in climate or nutrient supply can result in an apparent
constant average value below that of the potential maximum biomass.  Reichle
and his associates define this level as the maximum persistent biomass.
These ideas in different form were presented by  May [1973].

     We measure biomass of the various phases of the aquatic community
depending on how the phases are conceptually organized:  (1) trophic levels



                                MAXIMUM POTENTIAL BIOMASS
                              MAXIMUM PERSISTENT BIOMASS
Figure 12.   The effects of environmental fluctuations on steady state biomass
           in an ecosystem.  Three levels of persistent biomass are hypothe-
           sized based on nonvariant,  variant, and a highly variant environ-
           ment wherein severe perturbations occur [from Reichle et al.  1975],

(producer,  consumer,  decomposer), (2) organism level (as in Figure 11), and
(3)  photosynthesis/respiration (autotroph/heterotroph).  Note that indicator
organisms are assessed on the basis of  biomass or condition.

    Methods of biomass estimation include energy content, dry weight, ash
free dry weight, total or particulate organic carbon, chemical oxygen demand,
chlorophyll a, numbers, volume, and various measurements related to the above.
All  these methods offer differing levels of accuracy (in relation to "true"
ecosystem structure and function), precision, sensitivity and convenience.

    Biomass itself has an interpretation; as shown in Figure 11, biomass is
most meaningful in an ecosystem context when the individual organisms that
make up the community are considered.   Food chain and trophic level were
among the first concepts applied in attempting to understand communities.
It was soon realized that static (one-time) measures of biomass in specific
categories  were inadequate to understand communities.  Phenological and
successional biomass changes were important concepts that integrated time
with biomass and structure. Concepts related to production and consumption
(photosynthesis/respiration; autotrophy/heterotrophy), energy flow, and
nutrient cycling similarly integrated time with biomass and function.  There-
fore, we look at biomass as a measure of community  structure and function.
Instantaneous biomass provides a statement about structure while change in
biomass over an interval is a major aspect of function (productivity/
consumption/decomposition/toxicity). Generally, we use biomass to estimate

presence/absence, before/after, upstream/downstream, diversity, ..., but
rarely in the sense of Relohle et al.  [1975].  This latter concept depends
on carrying capacity (potential maximum biomass) definitions and we do not
yet know how to determine this value.

Indicator Organisms

     Although Patrick et al.  [1949] and Margalef [1951, 1975] have suggested
that community indices would be of more value than single indicator organ-
isms, there are still many stream studies oriented toward individual
organisms.  Taxonomic difficulties make indicator organisms less useful
[Cairns 1975, Bartsch and Ingram 1966].  Pielou [1975] argues convincingly
that the ecological value of a specific organism may be minimal or at best
difficult to assess.  Rare and endangered species are a legal exception
(Endangered Species Act. PL 92-205).  Broad measures of ecosystem function
would probably be unaffected by the presence or absence of rare and
endangered species; thus, procedures should consider these populations
separately.  Bioassays \_APHA 1976] are one method of developing criteria for
indicator organisms and for analyzing observed data [Sprague 1971].  Bioassays
have not been adequately related to community variables.  Physiological
indicators have been used to assess stress due to pollution [e.g.,  Lynch
1975] but again these variables are difficult to relate to community

Microbial Indicators--
     The system suggested by Kolkwitz and Marsson is based on indicator
organisms (both micro and macroorganisms) that define different water
quality conditions in streams reflecting self-purification (assimilation
capacity):  polysaprobic (gross pollution), alpha-mesosaprobic, beta-
mesosaprobic, oligosaprobic; also finer divisions are possible [Sladacek

     Many microbes have been suggested as useful in the Saprobien Index.  In
studies of laboratory systems Bick [1971] argues that ciliated protozoans
are useful for analyzing organic pollution.  Their rapid response and world-
wide distribution makes them suitable indicators; however, severe taxonomic
difficulties limit their use.

     Cherry et al. [1977] observed a significant increase (from 15-25 percent)
in percent chromogenic bacteria in cultures of bacteria one year after
chemical pollution was removed from a fast-flowing stream in Aiken, South
Carolina.  The stream (Tims Branch) had received  229,900  kg/yr of chemical
pollutants including mineral acids, caustics, salts, and organics (terchlo-
ethylene, trichlorethylene, and methanol).  The recovery level of chromogen
(25 percent) was comparable to a nearby control stream (28 percent).

     Curtis and Harrington [1971] found BOD and soluble organic carbon to be
the most reliable guides to the "slime-promoting properties of a water."
Other parameters measured were:  suspended solids, POit-P, NH3-N, Kjeldahl N,
N03, NOs, total carbohydrate.

     Curtis and Curds [1971] 'surveyed sewage fungus from 178 sites in


England, Wales, and Scotland.   Sphaerotilus natans and/or zoogleal  bacteria
were most often dominant.  Cluster analysis showed certain relationships
between organisms.   Diversity indices showed that slimes with mixed dominant
organisms also supported a greater variety of other organisms.  Ruthven and
Cairns [1973] studied effects of phenol, pH, and heavy metals on protozoans
in laboratory systems and found very sensitive and specific response to the
different metals relative to other organisms and individual protozoans.

     Lowe [1974] compiled, standardized, and centralized ecological informa-
tion concerning the requirements and pollution tolerance of fresh water
diatoms.  As had Patrick over the last 30 years, Lowe suggests that diatoms
(Bacillariophyta) are particularly useful as indicator organisms of the
condition of inland waters because they are relatively easy to identify to
the species level,  they are the most ubiquitous of aquatic organisms, and
each diatom species is thought to occupy a different niche in the aquatic
ecosystem with individual responses to chemical and physical  parameters.

     Three hundred  common species and varieties of fresh water diatoms are
tabulated with their environmental requirements and pollution tolerance.
The tolerance ranges for species are given for pH, nutrients, salinity
(holobion spectrum), organic (saprobien), current, general habitat, specific
habitat, seasonal distribution, temperature spectrum, and geographical
distribution and other information.  The data were compiled from 48 refe-
rences.  Tables of ecological  profiles make up 297 pages.

     Diokman [1969] found that differences in periphyton colonization of
agar coated slides  -- controls and containing suspected toxicant -- reflected
the nature of the toxicant being tested.  While interface conditions were
uncontrollable, this method was suggested as useful for screening potentially
introduced pollutants.

Aquatic Flora--
     Ray and White  [1976] suggest that because of accumulation processes,
vascular plants (Potamogeton and Equisetum] and the blue green alga,
Osoillatoria, be used to assess heavy metal pollution in aquatic systems.
The alga was found (with one exception) at the more polluted sampling sites
where the vascular plants were absent.  Some evidence was given indicating
that the alga accumulated lead and thus was a good indicator for lead
pollution.  The authors concluded that analysis of plant tissues, when
compared with control streams, is a reliable indicator of heavy metal

     Investigations of ecosystem change in Swedish lakes due to acid precipi-
tation provide insight to the subtler effects on aquatic communities \_Grahn
1970].  As these lakes have become acid, a succession has taken place in the
aquatic macrophyte community.  Sphagnum, which needs free COa (no HCOs) for
growth, is taking over littoral habitats formerly occupied by vascular
plants  (Isoetides,  Lobelia dortmanna, and Latorella uniflora].  In addition,
much of the microbial decomposer population had been destroyed, and decaying
plant matter was building up, further restricting higher plant habitat.

Aquatic Fauna--
     Eowm-Llle? and Scott [1977] discussed several  types of indices which are
distinguished by the kind of information they summarized.   They are critical
of diversity indices based on community structure  for not being sensitive
to physiological characteristics or ecological  affinities of the taxa in
the communities.  The authors also point out that  indices based on the
presence or abundance of one or more groups of organisms may be insensitive
to environmental change unless species-level analysis is done.  They
suggest an environmental index which considers both composition and ecology
of the community fauna as indicated by oligochaetes.

     By monitoring sediment selectivity and viability of a Great Lake amphi-
pod, Pontoporeia af'finis, Gannon and Beeton [1971] were able to evaluate the
potential impact of disposal of dredged harbor sediments.   The methods are
simple and inexpensive and can be adapted to other water pollution problems
and benthic organisms.

     Gaufin [1973] conducted 96 hr TLM's on 20 species of aquatic insects
and one amphipod to determine the effects of low DO,  elevated temperatures,
and low pH.  Longer term bioassays were also conducted which indicated
increased sensitivity with increased exposure.

     Beak [1977] compiled and standardized data on the environmental
requirements and pollution tolerance of 230 taxa of freshwater chironomids
in order to help in the evaluation of data from aquatic macroinvertebrate
collections taken for the assessment of water quality.  Chironomids were
selected because they can be identified to species level relatively easily
[Beok 1977].  However, larvae are the most common  benthic stage and are not
easy to identify.  They are among the most ubiquitous aquatic invertebrates,
and each taxon is thought to occupy a different niche in the aquatic
community.  Water quality spectral ranges for pH,  salinity, nutrients
(trophic status), degradable dissolved organics, oxygen, temperature,
turbidity, current, general habitat, specific habitat, seasonal distribu-
tions (emergence), feeding behavior, and geographic distribution are
summarized for each taxon according to developmental  stage.  The 28 years
of work by the authors provides most of the information.

     Rosenberg and Wiens [1976] used artificial substrates to assess the
significance of oil pollution on chironomids in Trail River in Northern
Canada.  Wilson and MoGill [1977] used the method  of collecting pupal skins
shed by members of the chironomid fauna which lodge and collect along stream
banks and vegetation to evaluate the water quality of an English stream
receiving treated sewage effluent.  The stream being  investigated could not
have been sampled by other conventional methods such  as kicknetting.
Despite the draw-back of only using one taxonomic  group, the authors were
able to show a definite relationship of species composition to water quality.
Chironomas riparius was the most responsive, being most dominant near the
sewage outfall, and less abundant in cleaner water.

     Pearson and Rosenberg [1976] studied the effects of organic matter
accumulation (cellulose) on the macrobenthos of fjordic systems in Scotland
and Sweden.  They considered that more generalized species occurred in the


benthic community as depth of aerobic sediment decreased.   The process of
deoxygenation of sediments is initially accompanied by an  increase in biomass
and productivity followed by an afaunal condition.   No species alone served
as a good indicator of community condition at any state of the succession.
Diversity indices (mathematical or graphical) were  advantageous, within
their limits, for comparing faunas at different stages in  the succession.
The use of indices must be qualified by presentation of specific data.

     Crowthev and Hynes [1977] studied the effect of road  deicing salt on
the drift of aquatic macroinvertebrates in Laurel Creek which flows through
urban Waterloo, Ontario.  This stream had maximum concentrations of Cl
(1770 mgA), Na (9550 mg/£), and Ca (4890 mg/£) during the winter of the
study period (1975).  Pulses of salt concentrations up to  800 mg/£ as Cl had
no effect on Hydropsyche betten-L, Chewnatopsyohe analis, and Gammarus
Pseudolirmaeus under laboratory trials.  Under field conditions, all
organisms drifted when concentrations of artificially added salt exceeded
1000 mg/5, as Cl.

     Anderson et al. [1978] used the fingernail clam (Musouliurn transversum]
as an indicator of water quality in the Mississippi and Illinois Rivers.  By
using the ciliary beating rate of the gills, they found this organism very
sensitive to heavy metals (6 * 10~" mg Zn/£) and unionized ammonia.  Investi-
gations of the effects of light, temperature, DO, sodium nitrate, sodium
sulfate, cyanide, Pb, Cu, Zn, suspensions of silica and illite clay, and
raw Illinois River water were made and they concluded that heavy metals and
unionized ammonia played major roles in removing the fingernail clam from
the rivers studied.

     A review of the literature dealing with the use of indicator organisms
to monitor trace metal pollution has been prepared by Phillips [1977] (189
references).  The author concludes from his review that despite the increas-
ing amount of information about metals in organisms, more attention must be
paid to the condition (weight, reproductive state,  etc.) and environment
(salinity, temperature, etc.) wherein the analyzed organism dwells.  Ignoring
these variables can lead to erroneous interpretation of data.  Bivalve
molluscs and macroalgae were the best studied organisms used as metal
pollution indicators.  It is recommended that both of these types of organ-
isms be used to give a more complete picture of the trace metal load.


     In considering community structure and function, steady-state and
requisite community stability were defined.  With correction to a seasonal
time frame, steady state exists only for a climax community.  A climax
community is stable by definition.  Stability exists under relatively
constant environmental conditions.  Although several properties of ecosystems
have been defined as representing stability, diversity and complexity of
niches in the community have not been accepted as measuring stability
\_0rians 1975].  May [1973] argued that complexity does not confer stability;
stability arises from coevolutionary development in ecosystems.  MaeMdhon et
al. [1978] derive similar arguments.  Stability increases with organization
of the community.  Thus measures of organization (negentropy) are measures


of community stability.

     Although acceptable measures of community stability are difficult to
define [Hulbert 1971], the concepts of community diversity have utility in
measuring stream ecosystem responses to perturbations.   Slobodkin and Sanders
[1969] assume that there is a connection between species diversity and the
physical properties (abiotic) of the environment.  They define a high
diversity category plus  three categories of low diversity (species poor)
environments:  (1) new environments (succession occurring); (2) severe
environments (predictable but with extreme environmental conditions such as
arctic seas, high salinity lakes, and hot springs.  These possess low
abiotic diversity); (3)  unpredictable environments (environmental conditions
vary extremely and frequently).  The authors advance arguments that environ-
mental predictability is a more important factor in controlling diversity
than other variables.   Thus diversity, by whatever measure [Pielou 1975,
Kaesler et al.  1978],  would be an adequate variable to assess impacts on
stream communities regardless of its relation to stability.
               and Sanders [1969] discussed the contribution of environmental
predictability to species diversity.  Many of their conclusions spell out
in ecological terms why reaches of streams impacted by urban runoff tend to
be species poor.   They point out that the combined effects of environmental
severity and unpredictability are such that a severe and unpredictable
environment, such as one might expect in a stream subject to the hydrology
and pollution from urban runoff, tended to be poorer in species than either
a severe or an unpredictable environment.  The authors also concluded that
generally different life history stages or organisms show different sensiti-
vity to environmental changes in "unpredictable" environments.  The authors
felt that competitive interactions are different in predictable and unpre-
dictable environments in such a way that fewer species persist in unpredict-
able environments.  Areas of low environmental predictability would be less
likely to be invaded by species from high predictability areas.

     Concepts of stability may be extremely dependent on "keystone" species
[Paine 1969] and/or "foundation" species [Dayton 1972].  Keystone refers to
a single species that controls community function and structure by its
presence.  Foundation species is the set of critical species that define the

     Other concepts related to stability refer to the ability of the commu-
nity to resist environmental changes (resistance) and the phenomenological
inverse of resistance, the ability of a community to return to a stable or
specific composition when stress (or change) is relieved (resilience).

     Stability, resistance, and resilience are properties of systems that
have been used to analyze the effects of environmental changes on aquatic
ecosystems.  A variety of diversity estimates have been used to estimate the
role of these concepts; diversity (the number of niches (species?; probably
deme) and individuals per niche in a community), niche richness (number of
niches), and evenness (distribution of individuals among the niches) have
been the most frequently cited.  In streams most of these estimates have
been applied to benthic invertebrates.


     Several  methods of calculating diversity have been devised.   Margatef
[1965] describes several approximating indices based  on varying levels  of
resolution in identifying and counting organisms;  also he estimated  diversity
based on visible light absorption ratios.

     Because of the lack of problem definition, recent discussions of
diversity and redundancy indices have not  clarified the issue of whether
diversity measures stability [Hamilton 1975,  Zand  1976, Logan 1977].   The
question of which is the correct method for estimating community stability
is probably unanswerable at this time.

     The question of which is the appropriate calculation for determining
information content in species counts of samples results from mathematical,
sampling, and interpretive differences.  According to Pielou [1975],  the
Shannon (Shannon-Wiener) index (#') is used for large communities where
sampling is used and the sampling does not reduce  the diversity \_Shannon and
Weaver 1949].  It is only an estimate.  The Brillouin index (#) is used for
collections (exact census of entire community).

     Because stream ecosystems are usually sampled, the Shannon index is
usually appropriate and the following approximation where all taxa are
assumed \_Simberloff 1978] known is used:

                 H, = - Y ifiog  if
                         .L .  N   ^e N

where n - number in taxon i, N = total number, and t - number of taxa.
Because of patchiness and similar clustering problems, sampling provides
significant error; the larger the sample,  the closer H" approaches the
"true diversity" (#')•  Hierarchical  diversity (H'} can be determined
depending on whether species, genera,  ..., or other taxa are used [Pielou
1975, Kaesler et al. 1978].

     Previously, redundancy was used in an incorrect manner and consequently
had little interpretive value for analyzing stream impacts [see Zand 1976].
Many investigators have used overlap, percentage similarity, and other
formulations that relate to the amount of common information (similar
species and/or numbers in each species) contained  in different sampling
locations (spatial) or at different times  (temporal).

     Haedriah [1975] tested the hypothesis that the Shannon species  diversity
and overlap (percentage similarity, PS) can be used together to assess
environmental quality.  Samples of fish populations in Massachusetts estua-
ries and embayments had Shannon diversity values (loge) that ranged  annually
from 0.4 to 2.4 in close correlation with pollutional stress (logic  popula-
tion, r - 0.84).  Little seasonal change in population composition as
indicated by high seasonal PS was found in areas of low annual diversity,
while areas of high diversity had lower seasonal PS.

     Brook [1977] compared a percentage community  similarity (PCS) index and
an index based on actual abundance of taxa in a study of zooplankton in Lake


LBJ, a reservoir in central  Texas.   He concluded that indices such as those
studied facilitate comparisons of control  areas and those receiving effluents.
He also concluded that an index based on actual abundance of taxa was too
sensitive to rare species, thus leading to high sensitivity to sampling
error.  A better result was  obtained using a PCS index indicating structural
functional differences between communities.   Brook recommended using
several indices that have different areas  of sensitivity.

     Although Hoautt [1975]  argues  otherwise, evenness, the ratio of observed
diversity to maximum diversity for  the observed number of species, appears
to be a valid concept and mathematical formulation.  Evenness should be
determined according to the  appropriate index, Shannon's or Brillouin's.
For Shannon's, assuming that the total number of species (5) are known,
evenness (£") is determined  as follows \_Pielou 1975]:

                               E' = ff'/log s

     Pratt and Coler [1976]  assume  they obtain a complete collection by
using artificial substrates  (rubble in baskets) and consequently they
suggest calculation of Brillouin's  diversity and evenness values to describe
effects of urban runoff on macroinvertebrates.  No data were presented.

     Previously, Wilhm [1970], Wilhn and Dorris [1968], Reger (1973],
Gislason [1971], and many others have used Shannon's diversity to describe
the effects of stream pollution.  Their results indicate that diversity
reflects pollution effects on stream communities.

     Stauffer et al. [1974]  found decreased diversity indices (Shannon) for
fish communities in areas of thermal discharge where temperature intervals
exceeded 80-87°F (26.7 - 30.6°C).  Also, fish collected from thermally
enriched areas were less "healthy"  than fish from colder waters.

     Stoneburner et al. [1976] compared the diversity estimates derived
from the Shannon-Weiner equation and the sequential comparison index (SCI)
developed by Cairns et al. [1968] and Cairns and Dickson [1971].  They
concluded through statistical comparison of the two indices that the
diversity estimates in terms H' and SCI are similar in their ability to
predict the effects of wastewater on aquatic community structure.  Also,
they concluded that diatoms  and macroinvertebrates show corresponding trends
of sensitivity to waste loading.

     Ghetti and Bonazzi [1977] compared indices of diversity (#')» species
richness (d), and evenness (e) with each other and with the Saprobien and
several biotic indices for the macroinvertebrate community, developing
equations for the conversion of one index  to another.  Diversity was the
most scattered variable.  Correlation of variables allowed definition of
four classes of water (non-polluted, weakly polluted, polluted, and strongly
polluted).  Also, their paper reflects the geographic variation in indices
used to assess polluted streams in  Europe  and the  USA.

     Kaesler et al.  [1978] redefine sampling so that the sample is a
"universe," i.e., a completely censused community  (collection).  Therefore,


they use Brillouin's diversity estimate (#).   The advantages are that it is
not biased and that it is more sensitive to low diversities.  Moreover, they
followed Pielou* s [1975] argument and showed  that no one actually uses the
Shannon diversity index (#') because the true species count (taxon or niche)
is unknown.  In most cases a comparison of events is desired and as long as
adequate attention to statistical matters is  maintained, the choice of indices
is immaterial.  P-ielou [1975] indicates that  the use of H as an estimator of
a large community that can only be sampled would be invalid.  Kaesler et al.
[1978] develop hierarchical diversities of trophic and functional morpholo-
gical  classifications that seem to hold promise.

     Other methods of reducing large quantities of species and count data
are being developed as ecological tools.  Boesak [1977] reviewed methods of
numerical classification of ecological  data.   Resemblance measures and
clustering methods are considered.  Guidelines for interpreting numerical
classifications and for applying ecological classifications are included.
The author encourages the use of numerical classification techniques, and
feels  that they can enhance interpretation of the classification data and
improve ecological insight.

     Smith and Greene [1976] describe numerical methods for describing marine
benthic invertebrate communities affected by  submarine outfalls.  Although
they suggest that at our present state of knowledge the methods are princi-
pally useful for generating hypotheses, they  do conclude that HaS concentra-
tion in sediment porewater has major impact on community composition.

     Kaesler et al. [1978] argue that species diversity is unnecessary;
generic level provides adequate information.   Moeller et al. [In press] are
using a stream classification system for assessing "community quality"
(diversity?) that relates to macroinvertebrate functional groups as defined
by the stream continuum theory [Cummins et al. 1973].  Heck [1976] argues
that only detailed community analysis will allow the interpretation of
impacts on aquatic communities.  However, the complexity of the system he
studied plus its incomplete physical description may be the reason for the
lack of demonstrated value of other community oriented measures.

     Based on the foregoing results, measurement of diversity as a means of
describing impacts of contaminants on stream communities is a valid method.
Generally, diversity should be interpreted as an operationally useful
measurement and no conclusions on its ecological significance should be
made.   Analysis of benthic invertebrate communities are preferred but other
groups (fish, periphyton) can provide useful  data.  Generic diversity would
generally be adequate and measurements should be comparative:  upstream/
downstream, before/after.  Calculation of related indices can provide
additional information.


     This section deals with those macroscopic properties that reflect the
fundamental processes of ecosystems:  (1) groups of organisms that manipulate
and transform significant amounts (> 10 percent) of energy and materials;
and (2) the turnover of those materials.  Perturbations of aquatic ecosystems


should be massive enough to affect these properties.   The perturbation does
not have to be massive enough to cause a permanent change but should provide
a measurable change that can be statistically demonstrated.   Moreover, the
change should last for a measurable period.

     We have discussed indicator organisms and diversity, both of which are
used frequently and with some success.  Although valuable concepts,  the
disadvantages and complexities of taxonomy and definition make them  less
useful.  Also the level of response is not quantitatively related to impact,
nor, at the present state of knowledge, is it possible to develop quantita-
tive relationships between impacts and the properties of functional  groups
and material turnover.  However, it seems more feasible to develop such
quantitative relationships for these properties than  previously were
developed for indicators and for diversity estimates  [e.g.,  Maguire  1971].

     The concept of functional groups began to evolve in earnest with Odum's
[1969, 1971, Rolling 1973] continued development of the trophic levels
concept (producer, consumer, decomposer) and energy flow through trophic
levels.  At the present time we use functional groups in the development of
concepts about material processing.  The foremost example of this approach
for freshwater stream ecosystems is the stream continuum theory which
describes the biogeochemistry of organic matter \Ulnshall 1967, 1968, Cummins
et al. 1973, Baling et al. 1975].  The role of functional groups of  inverte-
brates (chiefly aquatic insects) in processing organic matter depends on
the sources of energy and materials (natural or cultural allochthonous
organic matter; autochthonous organic matter), and the status of the environ-
ment itself (tree canopy, riparian vegetation, geology, etc.).  Generally,
changes in P/R ratios and types of functional groups  occur in relation to
each other and in relation to inputs of material and  energy.  Natural
undisturbed communities in a given environment might  have P/R < 1.0  while
after disturbance, the same environment might shift to P/R > 1.0.  The
consequences would be to increase the proportion of specific functional
groups (grazers) relative to the natural condition (shredders, ...).

     Further work by Moeller et al. [In press] traces dissolved organic
carbon (DOC) transport through streams of differing morphology.  Although
they experienced difficulty in relating dissolved organic carbon to  parti-
culate organic carbon (POC), they were able to explain essentially all of
the DOC variation by discharge (flowrate), watershed  area, and link  magni-
tude (a measure of stream order).  Low DOC watershed  output came from
watersheds relatively undisturbed by human activities while higher values
often were associated with culturally influenced larger streams.  However,
the relationship of DOC (or POC) to stream impacts of urban NPS runoff is
not definable at this time.

     Riohey et al.  [1978] argue that one means of assimilating these
specific processes into a simpler broad scope hypothesis is to consider
carbon flow.  By determining the total carbon flow in a lake ecosystem (CT]
and the portion of the flow that cycles (cc], they define a cycling  index
(Cz) that could be related to lake condition (Ci = CC/CT] with further
information and analyses.  The method of determining  turnover time depends
on estimating biomass in elemental equivalents and then estimating the


elemental flow through the biota using isotopic tracers or appropriate
chemical analysis.   The overall  concept is related to that of Reiohle et al.
[1975] who summarize turnover times for terrestrial  systems (initial  mass/
change in mass per time) for C,  N, and Ca (Table 9)  and energy dynamics
(which they term ecosystem metabolism) (Table 10).  Stwnm and Baooin-L [1978]
show that environmental disruption in lakes results  in less energy efficiency,
more rapid biogeochemical  cycling of elements, increased productivity, and
loss of diversity.   All of these methods of analysis tend to reduce ecosys-
tems to simpler terms, masking those properties that make them unique from
each other.  This is a necessary step in developing  an encompassing theory
of ecosystem operation.  For the purposes of assessing impacts such as urban
NFS runoff impacts on aquatic communities, this approach seems to be the
most valid.  Unfortunately, there are not data that  allow assessment of
this hypothesis.

     Other approaches have included the use of microecosystems (and of models
as previously discussed in this  chapter).  Microecosystems permit the study
of communities under defined and controlled conditions wherein material
turnover and energy flow can be  assessed relative to specific perturbations
\PoTce11a et al. 1975, Mitchell  and Buzzell 1971, Cooke 1971, W-itt et al.
1979, Geisy 1978].   It is  possible that these systems present the best
approach for identifying those processes and hypotheses that would be
testable under field conditions  where random input,  varying magnitude
impulse loadings cause changes in stream community function and structure.

     Cornell et al. [1976] have  developed some concepts of change in diver-
sity, biomass, and evenness over time that seem to show promise as a
sensitive analytical means of assessing community structural response.  The
problem with applying diversity or biomass directly to assessing environ-
mental problems is that those indices have not been related to changes in
time except comparatively.  Cornell et al. argue that their concepts
integrate temporal patterns.


     A minimal set of response variables that relate to assessing impacts of
contaminants on aquatic communities would include habitat description,
generic diversity of aquatic macroinvertebrates, an assessment of carbon
turnover rates and a minimal set of water quality variables that relate to
the major contaminant inputs (DO, nutrients, metals, suspended solids,
salts).  Carbon is the major ecological element and stream macroinvertebrates
process the major quantities of materials and energy.  The turnover of
materials such as carbon would likely be affected by variations in water
quality caused by contaminant input.  Some inventory studies should be
made to assess populations of rare and endangered or other significant
species where appropriate.

     Sampling in time and space should be adequate for comparative purposes
(before/after; upstream/downstream).  If an understanding of ecosystems
were adequate to predict potential energy and material flow in communities,
such comparisons would  be unnecessary.


          REICHLE ET AL.  1975]1
Temperate Deciduous
  Forest  Component
                                      Turnover time, years
Forest Biomass
< 5
< 5
1 Comparison based on assessment of many variables; see Reiohle et al.  [1975]
  for explanation and footnotes.

           ET AL.  1975]2
                                          g Carbon/nfyr
Gross primary production (GPP)      1620       1320
Autotrophic respiration (RA)         940        680
Net primary production (NPP)         680        600
Heterotrophic respiration (RH)       520        370
Net ecosystem production (NEP)       160        280
Ecosystem respiration (RE)          1470       1050
Production efficiency (RA/GPP)
Effective production (NPP/GPP)
Maintenance efficiency (RA/NPP)
Respiration allocation (RH/RA)
Ecosystem productivity (NEP/GPP)
  Comparison based on assessment of many variables;  see Reiohle et al.  [1975]
  for explanation and footnotes.

                                SECTION VI



     The selection of case  studies  was not  intended to be exhaustive nor
restricted to urban runoff  events.  Generally,  the approach used was to look
for stream impact  analyses  and  use  those which  were illustrative of the
principles outlined previously.   Major reasons  for this approach were the
lack of studies related  to  urban  runoff impacts on, and the very few related
to, impulse inputs to stream ecosystems.


     Harrell [1978] analyzed the  diversity  effects of flooding on the Devil's
River in Texas.  A fortuitous flooding event  (ninth largest flood on record)
occurred midway during Barren's  study to assess  the fish community of the
river.  Marked changes in habitat occurred  as a result of the flooding;
distinctness of spring,  riffle,  pool,  channel,  and intermediate habitats
generally were blurred and  riffle-like habitat-types were more frequent.
Fish community diversity declined.  Harrell suggests that the ecological
plasticity of the  communities allowed  shifts  in the roles and associations
of fish species and thus allowed  the community  to be maintained.  In this
respect, the community is stable; naturally stressful environments \_siobodkin
and Sanders 1969,  Or-Lans 1975]  would require  dominance and maintenance of a
few species.


     Although many oil spill impact studies on  marine ecosystems have been
done [e.g., Foster and Holmes 1977, Nelson-Smith  1977, Dicks 1977], very
few similar types  of studies have been done in  freshwater systems.  However,
spill impacts are random impulse  type  events  and  therefore are analogous to
urban NPS runoff impacts on streams, although  they occur less frequently.

     Cairns and Diokson  [1977]  describe two stream systems that were impacted
in this manner, the Clinch  River and the Roanoke  River.  On June 10, 1967,
caustic wastewaters (pH  = 12) from  a fly ash  pond associated with a power
plant adjacent to the Clinch River  were released  through a collapsed dike
and caused extensive damage to the  aquatic  community.  The inflow was 40
percent of the normal Clinch River  flow at  that time \_Cairns et al. 1970].
In 1971 an acid spill occurred.   Meanwhile  during this period, day-to-day
operations of the power  plant and periodic  flooding had short term effects
on the aquatic biota.  The  authors  and co-workers assessed benthic


macroinvertebrates at 21  ecologically similar stations during the years 1969,
1970, and 1971 and analyzed that data based on numbers, biomass, diversity,
and presence/absence (cluster analysis).

     Approximately 200,000 fish were killed by the spill; however, most of
the analysis centered on  benthic invertebrates because invertebrates:

1.   are relatively sessile and cannot avoid stress;
2.   have long and complex life cycles;
3.   are important members of the food web and affect related organisms;
4.   have sampling techniques that are more reliable than for other organisms;
5.   have more biological information obtained per dollar invested.

     The initial  measurements (June 20, 1967) were made by State of Virginia
employees and Cairns and  associates analyzed these data to show that the
number of taxa and organisms per unit area were severely impacted by the
Station No.        Description               RM    Taxa      Benthos/ft2
     1           Upstream (reference)      +21.0    24           520
     2           Upstream but backwater    ,  n c     0            on
                   effects                 + U'b     9            30
     3           Downstream                -  0.3     0             0
     4           Downstream                  3.5     5           270
     5           Downstream                  9.5    11           275
     6           Downstream                 12.5    15           650
     7           Downstream                 17.2     8           640
     8           Downstream                 30.5    11           175
     9           Downstream                 40.2    22           750
    10           Downstream                 56.5    18            80
    11           Downstream                 65.5    22           260
    12           Downstream                 77.5    32          1070

     The authors  noted that the benthic community was 8 percent midges at
Station 1 but increased to more than 60 percent at damaged stations, illus-
trating their tolerance to stress.  By Station 10, the midges were less than
30 percent.  The  pattern  at these lower stations was that of lesser impact
(due to dilution  and physico-chemical reactions).

     Recovery was not rapid and still was an  ongoing process for the sampling
collected in July-December, 1969, for the same and additional stations:
                                                Total          Diversity
Station No.   Description   RM     Taxa      Benthos/ft2      (H", log 2)

   1-4        Above the   + 1.5   48-54      24.6-145.2        2.97-4.03
              spill ( =4)  to
              range       +26.5
    7                     - 0.2     43           47.0          2.92-4.21
    8                     - 0.9     33            2.4          3.64-4.35
    9                     - 3.9     36            4.0          2.88-4.01
   10                     - 9.5     42            8.0          3.19-4.06
   11                     -18.5     46           18.8          3.61-4.17


     The biotic density was considerably depressed but diversity had reached
control levels.  Unfortunately no data on diversity were presented for the
1967 data so it was not possible to analyze diversity as a function of time
according to Cornell et al.  [1976].  The longer life cycle and non-motile
organisms were those that had not returned; thus, recovery is a function of
recolonization time which is controlled by life cycle and motility.


     Ball et al.  [1973] analyzed three streams in Michigan that illustrate
the impacts of pollutants on stream communities:  (1) the Jordan River was
least developed with the downstream site differing from the upstream site
because of a fish hatchery effluent; (2) the AuSable River was affected
chiefly by recreational use but also by a previously thriving lumber
industry and forest fires; (3) the Red Cedar River was in a more urban/
industrial region and, although having received waste effluents directly in
past years, urban runoff and some industrial inputs were more common.
Generally upstream stations were better quality than downstream.

     The authors and their colleagues had performed a variety of measurements
that they related to the rivers as affected by material inputs.  Although
they concluded that the rivers were geochemically similar, the slope of the
Red Cedar River was not as steep as the other two.  Generally, total organic
carbon, non-carbonate hardness and total phosphorus were greater at the lower
station of the Red Cedar River than the other stations.  Suspended matter
was considerably greater for both stations on the Red Cedar River than for
the other rivers.  Generally toxins (metals, pesticides, and PCB), inorganic
nitrogen compounds and chloride were greater on the Red Cedar.  Although
data on macrophytes and fish were discussed, those data were not directly
comparable to other measures because species and their biomass were not
always present at all the stations.

     The most important variables of comparison are listed in Tables 11 and
12.  The AuSable River seemed most productive based on diurnal DO fluctua-
tions although the gross productivity of the Red Cedar River was greater.  A
comparison of average P/R ratios indicated that the better quality rivers
(Jordan and AuSable) had fairly typical values for clean midwestern rivers
while the Red Cedar River was the most variable and had values outside of
the normal range of P/R ratios.  This may have been associated with the
considerable input of organic matter in the lower Red Cedar, possibly from
urban runoff.

     The effects on the community structure were considerable as measured
by diversity.  Although diatom diversity was greater for the Red Cedar River
than the AuSable, and for invertebrates the AuSable was greater than those
of both the other rivers.  Also, the error associated with the estimated
diversity for the AuSable and the Red Cedar Rivers was greater than the
Jordan.  This observation relates closely to the theoretical development
of A#' as proposed by Cornell et al. [1976].  Based on these data  (Tables 11
and 12), one would conclude that any of the variables, DO, diversity
(periphyton, invertebrates), and productivity would be of value in assessing
urban NPS runoff impacts in freshwater streams.


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     A variety of studies of urban NPS runoff impacts on stream communities
have been performed and some will  be summarized herein but only one will  be
discussed in detail as exemplifying the processes involved.

     Chisholm and Downs [1978] studied the benthic invertebrate ecology of a
small West Virginia stream which received sediment from construction of a
super-highway.  In comparison with a nearby control  stream,  the diversity
index, generic count, and total  count of invertebrates indicated severe
reduction or destruction of the  benthos in the impacted stream.  The greatest
degradation occurred in areas of highest sediment movement.   When construction
ceased, the benthic population of the impacted stream recovered to a
comparable level with the control  stream.  Colonization was  from upstream,
unimpacted tributaries.

     Except in streams subject to combined sewer and sanitary overflow
discharges, Ragan and Dieteman [1975] found streams  in areas of Maryland,
urbanized during and after the 1950's, to be of good quality.  This was not
expected but the authors attributed this to the lack of sanitary discharges
in the area, and to the large (>10 square miles) size of the drainage basins
being studied.  Even though traditional water quality parameters did not
show dramatic changes, pollution sensitive fish species (e.g., Rosyside Dace)
were no longer found in urbanized areas.  A nearby,  unurbanized stream had
the same 21 species of fish that were present in 1912.

     DiGiano et al. [1975, 1976] directed their study in Greenfield, Massa-
chusetts, toward both short term impacts on water quality and the longer
term disruptions of the benthic  macroinvertebrate population caused by urban
runoff.  Using weekly grab samples from the Green River, they found that
total P, turbidity, chloride, and total coliform concentrations increased
with runoff events.  They were unable to show conclusively that BOD, TOC,
and oil and grease concentrations increased with runoff.  It was suggested
that more detailed analyses of both discharge and concentration be used in
determining the mass loading for relatively short (3.2 mile) river reaches.
Distinguishing between combined  sewer overflows and  stormwater runoff events
is also important, particularly  for bacteriological  studies.  In an analysis
of three specific storm events sampled at 15 or 30 minute intervals, they
found considerable variations from storm to storm of total pollutant loading
(Table 13).

     The study of the benthic diversity of the macroinvertebrate community
(Figure 13) of the stream showed constancy in diversity (Brillouin's) and
composition of central collections.  However, downstream collection, in
urbanized areas of the stream, showed reduced diversity and a community
shift to more pollution tolerant (polysaprobic) species.  In addition, the
seasonal variation in diversity  increased at downstream stations.  The
physical character of the stream bottom and availability of substrate
deteriorated due to sedimentation.  Sedimentation did not explain all of the
changes in benthos.  The contribution of contaminants from natural and man-
made sources remains to be determined.  Bioaccumulation of metals may result
through the detritus-based food  chain.  The authors  stated that stress from



Parameters          March 24, 1975      June 12, 1975      July 9, 1975

BOD (Ib/day)        -95.0 - 6100        1180   - 2550        22.5 -  380
Total P  (Ib/day)          -               34.6 -   83.6       1.0 -   21.5

TOC (Ib/day)      -1040   - 8040        1160   - 9900        95   - 2570
Chloride (Ib/day) 10400   - 12500       5240   - 7830      2130   - 5210
these heavy metals on benthic  invertebrates  needs  more investigation.


     The major conclusions about these case  histories  are:

1.    macroinvertebrates and periphyton seem  to be  excellent for monitoring
     impulse impacts on streams;
2.    diversity and biomass are often used for the  analysis  of monitoring but
     they are not always adequately sensitive;
3.    indicator organisms seem  not to be useful;
4.    water quality variables do not always reflect stream impacts directly.
     The biotic community seems to be more important in reflecting impulse
     inputs of contaminants.

     A scenario to illustrate  the hydrologic/physico-chemical variables and
biological/ecological variables for determining urban  MPS runoff impacts on
stream ecosystems would help illustrate some of the problems inherent in
analyzing these kinds of systems.  A storm occurs  at a specific time for a
specific interval.  The precipitation begins to run off depending on antece-
dent storms, soil moisture and impervious surfaces.  The pickup of contami-
nants varies with antecedent "washings," velocity  of runoff and natural and
cultural activities in the runoff basin.  Transport of the  water and contami-
nant materials enters the stream diffusely or at a point or several  points
and over an interval and pattern controlled  by the transport system. This
pattern provides the random event/magnitude  impulse of contaminants.

     The impulse of contaminants is diluted  within the stream and has varia-
ble impact depending on peak concentration of contaminants, duration of
exposure and life cycle period of the organisms making up the community.
Survival of specific organisms/taxa/niche is controlled by such events.
Eliminated community components can be replaced by recolonization but the
interval depends on out-of-stream or upstream sources  and physiological
variables (length of reproductive cycle, motility).  Other niches may be
opened by such inputs but generally communities become simpler, nutrient
cycling speeds up (shorter turnover times),  and efficiency decreases.   The
variables that seem most useful in reflecting these processes are not
necessarily the classical biotic density and diversity measurements used



                                                      	• STATION I

                                                            STATION I

                                                            STATION HI
                  STATION I

                  STATION IZ
                                                         •-« STATION I

                                                         -* STATION 3Z
                                                            STATION I

                                                            STATION 21
                  MAR.I-MAY3I  JUNEI-AUG.3I
                   1974          1974
                                COLLECTION PERIOD
Figure 13.   Comparison of  the species diversities of  the  benthic macroinver-
             tebrates between Stations 1  and 6, Green  River, Greenfield,
             Massachusetts  \_Di-Giano et al.  1976].
previously  but those processes associated with community function:
flow and  material cycling.

                              SECTION  VII


                    IMPACTS  ON STREAM  ECOSYSTEMS

     Throughout this report,  we  have  asserted  that  random  event and magni-
tude impulses to stream  ecosystems  would result  in  changes in  ecological
variables related to community structure and function.   In this section, we
propose an approach for  assessing such  impacts.   Because the spatial  distri-
bution of the individuals of  a species  frequently follows  a log-normal
distribution \_Slooomb and Diokson 1978]  and the  probability distribution of
runoff events can be determined, a  stochastic  approach  suggests itself.  A
stochastic model is a probabilistic model  that includes  a  time dependent
component.  Some techniques that illustrate the  application of stochastic
models to water quality  data  indicate the feasibility of this  approach.
Krishman et at. [1974],  Padgett  et  al.  [1977], and  Padirtanabhan and Delleur
[1978] apply this approach to BOD modeling in  streams and  thus are able to
calculate the mean DO in a stream,  the  error about  the  mean, and  the
probability of violating a specific criterion  or standard. All three
generated data using storm runoff models because actual  data were insuffi-

     Smith [1978], in a  recent review of optimization in ecological systems,
discusses the concept of attempting to explain ecosystem functional varia-
bles using optimization  models while  listing recent attempts to model  and
optimize ecological processes.  An  example of  how temperature  modeling can
be used for assessing catastrophic  events such as climatic and flow varia-
tions that result in extreme  water  temperatures  illustrates a  reasonable
duplication of actual  data [Morse 1978].  It is  simple  to  see  how such events
could result in an ecological catastrophe but  it is difficult  to  assess the
magnitude of the ecological response  and even  more  difficult at this  time
to develop the probabilistic  relationships between  the  impulse and the
response.  For example,  Slatkin [1978]  used a  model study  to show that
stability existed only when the  time  scale for environmental change is
roughly comparable to the average response time  of  those changes. Means
and variances did not determine  extinction tendencies.


     Impulse loadings can vary seasonally with material  (nutrient, toxicant,
oxygen demand) or energy (allochthonous  reduced  carbon,  light, temperature)
inputs.  Such random impulses are exemplified  for a hypothetical  seasonal
cycle for fish in Figure 14.   Two sensitive stages  are  illustrated:   (1) egg


               o  I
               _  T3






it Q


                     . Q.

                                  0) r—
                                  -C J3
                                  •!->  1X5
                                  O ••-
                                  CX S-
                                  >> fO
                          3H8VmVA  JO


growth requirement for dissolved oxygen (the DO is utilized by degradation
of organic matter in urban NFS runoff); (2) fry development requires lower
streamflow to catch food and maintain safety from predation by large fish.
If events cause changes in DO and flow variables to violate the requirement
for a sufficient time and a sufficient magnitude, that year class could be
eliminated.  Continued annual cycles where violations occur could eliminate
the species.

     Thus the probability that a species or individual will not survive is
equal to the probability that at least one event will occur that will
eliminate the species or individual.  For example, where flow is the event
that threatens the species:
     prob (species eliminated) = 1  - prob (#<


Interval for

Imput of Urban

Runoff Carried


                          TIME ,  YEARS

 Figure 15.  Percent survival  decreases with more severe urban  runoff  events,

                                                           Space  B
                                                  Space A

 Figure 16.  There is experimental  error for system variables which  adds  to
             the variance caused  by response to stochastic inputs.

     Ecological  variables for assessing these impacts would be measurements
of structurally and functionally defined biomass, photosynthesis, respira-
tion, diversity, in time and space.   The materials, carbon and oxygen, would
be measured.  Values would be estimated using these units and energy units.
Water quality and material input measurements should be made, especially
throughout the flood hydrograph.  The number of required measurements could
be determined by using rarefaction procedures \_simberloff 1978].   Eventually,
based on a probability defined by policy, management or law, the stream
sampling in time and space could be  defined on the basis of the ecological
resource being protected.

                               SECTION VIII


Anderson, K. B., R.  E.  Sparks,  and A.  A.  Paparo.   1978.   Rapid assessment of
     water- quality using the fingernail clam Musculium transversum.   UILU-
     WRC-78-0133.   Wat.  Res. Cen., Univ.  of Illinois  at Urbana-Charipaign,
     Urbana, IL.  115 p.

APHA.  1976.  Standard methods.   APHA.  New York,  NY.   pp.  685-882.

Ball, R. C., N.  R. Kevern,  and  T.  A.  Haines.   1973.   An ecological  evaluation
     of stream eutrophiaation.   Tech.  Rpt.  #36.   Inst.  Wat.  Res., Mich.  St.
     Univ.  East Lansing, MI.

Barkdoll, M. P., D.  E.  Overton,  and R.  P.  Betson.   1977.   Some effects  of
     dustfall on urban stormwater quality.   Journal WPCF 49:1976-1984.

Bartsch,A.  F.,and W. M. Ingram.   1966.   Biological analysis of water
     pollution in North America.   Verh. Internat.  Verein.  Limnol. 16:786-800.

Beck, W., Jr.  1977.   Environmental requirements and  pollution tolerance of
     common freshwater chironomidae.   EPA-600/4-77-024.   U.  S. EPA,  Cincinna-
     ti, OH.  261  p.

Bick, H.  1971.   The potentials  of ciliated protozoa  in the biological
     assessment of water pollution levels.   In:  Proc.  Int.  Symp. on
     Identification and Measurement of Environment Pollutants.  Ottawa,
     Ont., Can.   pp.  305-309.

Boesch, D. F.  1977.   Application of numerical classifications in ecological
     investigation of water pollution.  EPA-600/3-77-033.   U.  S.  EPA,
     Con/all is,  OR.   115 p.

Boling, R. H., Jr.,  E.  D. Goodman, J.  A.  Van Sickle,  J.  0.  Zimmer,  K. W.
     Cummins, R. 0.  Petersen, and S.  R. Reice.  1975.   Toward a model of
     detritus processing in a woodland  stream.  Ecol.  56:141-151.

Bovee, K. D., and R.  Milhouse.   1978.   Hydraulic simulation in instream flow
     studies:  Theory and technique.   IFI  Paper No. 5,  FWS/OBS-78/33.
     U. S. FWS,  Ft.  Collins, CO  80526.  131  p.

Bradford, W. L.   1977.   Urban stormwater  pollutant loadings:  A statistical
     summary through 1972.   Journal WPCF  49:613-622.


Brock, D. A.   1977.   Comparison of community similarity indexes.  Journal
     WPCF 49:2488-2493.

Brownlee, R.  C., T.  A.  Austin, and D.  M.  Wells.   1970.   Variation of urban
     runoff with duration and intensity of storms.   Interim report.   Wat.
     Res. Cen., Tex. Tech Univ.  Lubbock, TX.   67 p.

Bryan, E. H.   1974.   Concentrations of lead in urban  stormwater.  Journal
     WPCF 46:2419-2421.

Cairns, J., Jr.  1975.   Quantification of biological  integrity,  in:  The
     Integrity of Water.   U.  S. EPA.  Sup. of Doc.   055-001-01068-1.  Wash.
     D. C.  20402.   pp.  171-188.

Cairns, J., Jr., D.  W.  Albaugh, F. Busey, and M. D.  Chariay.  1968.  The
     sequential comparison index:  A simplified method  for non-biologists to
     estimate relative  differences in  biological diversity in stream pollu-
     tion studies.   Journal WPCF 40:1607-1613.

Cairns, J., Jr., and K.  L. Dickson.  1971.  A simple  method for the biologi-
     cal assessment of  the effects of waste discharges  on aquatic bottom-
     dwelling organisms.   Journal WPCF 43:755-772.

Cairns, J., Jr., and K.  L. Dickson.  1977.  Recovery  of streams from spills
     of hazardous materials.   In:  Recovery and Restoration of Damaged
     Ecosystems.  J. Cairns,  Jr., K. L. Dickson, and  E. E. Herricks, eds.
     Charlottesville, VA.  pp. 24-42.

Cairns, J., Jr., K.  L.  Dickson, and J. S. Grossman.   1970.  The biological
     recovery of the Clinch River following a fly ash pond spill.  25th
     Industrial Waste Conference.  Part One.  pp. 182-192.

Cech,  I., and K. Assaf.   1976.  Quantitative assessment of changes in urban
     runoff.   ASCE J. Irrig.  Drain. Div.  102:119-125.

Chen,  C. W.,  and R.  F.  Selleck.  1969.  A kinetic model for fish toxicity
     threshold.  Journal WPCF 4-1:294-308.

Cherkauer, D. S.  1977.   Effects of urban lakes on surface runoff and water
     quality.  Wat.  Res.  Bull.  13:1057-1067.

Cherry, D. S., R. K. Guthrie, F. L. Singleton, and R. S.. Harvey.  1977.
     Recovery of aquatic bacterial populations in a stream after cessation
     of chemical pollution.  Water, Air,  and Soil Poll. 7:95-101.

Chisholm, J.  L., and S.  C. Downs.  1978.   Stress and recovery of aquatic
     organisms as related to highway construction along Turtle  Creek, Boone
     County, West Virginia.  Geol. Sur. Wat.-Supply Paper 2055.  U. S. GPO,
     Wash., D. C.   20402.  40 p.

Christensen, E.  K.,  J.  Scherfig, and M.  Koide.   1978.   Metals from urban
     runoff in dated sediment of a very shallow estuary.   Environ.  Soi.
     Tech.  12:1168-1173.

Cooke, G. D.  1971.   Aquatic laboratory microsystems and  communities.  In:
     The structure  and function of fresh-water microbial  communities.  John
     Cairns, Jr.  ed. Amer.  Micro. Soc. Symp.  pp. 47-85.

Cordery, I.  1977.   Quality characterization  of urban storm water in Sydney,
     Australia.   Wat.  Res.  Res. 13:197-202.

Cornell, H., L.  E.  Hurd,  and V. A. Lotrich.  1976.  A measure of response
     to perturbation used to assess structural  change in  some polluted and
     unpolluted stream fish communities.  Oecologia (Bert.) 23:335-342.

Crowther, R. A.,  and H. B.  N. Hynes.  1977.  The effect of road deicing salt
     on the drift of stream benthos.  Environ.  Pollut. 14:113-126.

Cummins, K. W.,  R.  C.  Petersen, R. 0. Howard, J. B. Wuycheck, and V. I.  Holt.
     1973.   The utilization of leaf litter by stream detritivores.   Ecol.

Curtis, E.  J. C., and C.  R. Curds.  1971.  Sewage fungus  in rivers in the
     United Kingdom:  The slime community and its constituent organisms.
     Wat. Res. 5:1147-1159.

Curtis, E.  J. C., and D.  W. Harrington.   1971.   The occurrence of sewage
     fungus in rivers in the United Kingdom.   Wat. Res. 5:281-290.

Dale, J. T.  1978.   Bottling rainstorms—Chicago1 s tunnel and reservoir plan.
     Journal WPCF 50:1888-1892.

Davis, E. M., D.  M.  Casserly, and J. D.  Moore.   1977.  Bacterial relation-
     ships in stormwaters.   Wat. Res. Bull. 13:895-905.

Davis, J. A., and J. 0. Leckie.  1978.  Effects of adsorbed complexing
     liquids on trace metal uptake by hydrous oxides.  Env. Soi. and Tech.

Dayton, P.  K.  1972.  Toward an understanding of community resilience and
     the potential  effects  of enrichments to  the benthos  of McMurdo sound,
     Antarctica.   In:   Conservation Problems  in Antarctica,  pp. 81-95.

Dickman, M.  1969.   A quantitative method for assessing the toxic effects of
     some water soluble substances based on changes in periphyton community
     structure.   Wat.  Res.  3:963-972.

Dicks, Brian.  1977.  Changes in the vegetation of an oiled Southampton
     water salt marsh.   In:  Recovery and Restoration of Damaged Ecosystems.
     J. Cairns,  Jr., K. L.  Dickson, and E. E. Herricks, eds.  Univ. Press of
     Virginia.  Charlottesville, VA.  pp. 208-240.


DiGiano, F.  A., R.  A.  Coler, R.  C.  Dahiga, and B. B.  Berger.  1975.  A
     projection of pollutional  effects of urban runoff in the Green River,
     Massachusetts.   In:   Urbanization and Water Quality Control.   Annual
     Symp.  of the Amer.  Wat. Res.  Assoc.  Proc. 20.  pp. 28-37.

DiGiano, F.  A., R.  A.  Coler, R.  Dahiga, and B. B. Berger.  1976.  Characteri-
     sation of urban runoff - Greenfield, Massachusetts.  Phase II.  Wat.
     Res.  Res.  Cent.,  Univ.  of Mass.   Amherst, MA 137 p.

Dillon, P.  J.,  and F.  H.  Rigler.   1975.  A simple method for predicting the
     capacity of a lake  for development based on lake trophic status.  J.
     Fish.  Res. Board.  Canada 32:1519-1531.

Diskin, M.  H.,  S. Ince,  and K.  Oben-Hyarko.  1978.  Parallel cascades model
     for urban  watersheds.   ASCE J.  Hydraul.  Div. 104:261-276.

Doehring,  D. 0., and M.  E.  Smith.   1978.   Modeling the dynamic response of
     floodplains to urbanization in eastern New England.  Environ. Res. Cen.
     CSU,  Ft. Collins,  CO  80523.   95 p.

Eckenfelder, W. W., and  D.  L. Ford.   1970.  Water pollution control:  Experi-
     mental procedures for process design.  Pemberton Press.  Austin and
     New York.   269 p.

Ellis, J.  B.  1976.  Sediments and water quality of urban storm water.  Wat.
     Serv.   80:730-734.

Emery, R.  M., C. E. Moon, and E.  B.  Welch.  1973.  Enriching effects of urban
     runoff on  the productivity of a mesotrophic lake.  Wat. Res.  7:1505-

Eppley, R.  W.,  and W.  H.  Thomas.   1969.  Comparison of half-saturation
     constants  for growth and nitrate uptake of marine phytoplankton.  J.
     Phycol. 5:375-379.

Field, R.   1978.  Discussion:  Effects of storm frequency on pollution from
     urban runoff.   Journal WPCF 50:974-975.

Field, R.,  and R. Bowden.  1976.   Urban runoff and combined sewer overflow.
     Journal WPCF 48:1191-1206.

Field, R.,  R. Bowden, and K. Roygonyi.  1977.  Urban runoff and combined
     sewer overflow.  Journal WPCF 49:1095-1104.

Field, R.,  and B. B. Gardner.  1978.  Urban runoff and combined sewer over-
     flow.   Journal WPCF 50:1170-1185.

Field, R.,  and D. Knowles.  1975.  Urban runoff and combined sewer overflow.
     Journal WPCF 47:1352.

Field, R.,  and P. J. Szeeley.  1974.  Urban runoff and combined sewer over-
     flow.   Journal WPCF 46:1209-1226.


Field, R., and P.  Weigel.   1973.   Urban runoff and combined sewer overflow.
     Journal WPCF 45:1108-1115.

Fitzgerald, G. P.   1969.   Some factors in the competition or antagonism
     among bacteria, algae, and  aquatic weeds.  Phycol.  5:351-359.

Foster, M. S., and R. W.  Holmes.   1977.  The Santa Barbara oil spill:  An
     ecological disaster?   In:  Recovery and Restoration of Damaged Eco-
     systems.   J.  Cairns,  Jr., K. L. Dickson, and E. E. Herricks, eds.   Univ.
     Press of Virginia.  Charlottesville, VA.  pp. 166-190.

Gakstatter, J. H., T. E.  Maloney, and F. B. Lotspeich.  1977.  Guidelines
     for assessing the benefits  of best management practices to stream
     ecosystems.   EPA CERL-039.   Corvallis, OR 97330.  19 p.

Gannon, J. E., and A. M.  Beeton.   1971.  Procedures for determining the
     effects of dredged sediments on biota--Benthos viability and sediment
     selectivity tests. Journal WPCF 43:392-398.

Gates, C.  D.  1975.   Urban runoff in Binghampton, New York.  In:  Urbaniza-
     tion and Water Quality Control.  Annual Symp. of the Amer. Wat. Res
     Assoc., Proc. 20.  pp. 38-44.

Gaufin, A. R.   1973.  Water quality requirements of aquatic insects.  Office
     of Research and Monitoring.   EPA-600/3-73-004.

Geisy, J.  P.  1978.   Microcosms  in ecological research.  SREL.  6th Annual
     Symp., Aiken, SC.  In preparation.

Geldreich, E.  E.,  L. C. Best, B.  A. Kenner, and D. J. Van Donsel.  1968.
     The bacteriological  aspects of stormwater pollution.  Journal WPCF

Ghetti, P. F., and G. Bonazzi.  1977.  A comparison between various criteria
     for the interpretation of biological data in the analysis of the
     quality of running waters.   Wat. Res.   11:819-832.

Gislason,  J. C.  1971.  Species  diversity of benthic macroinvertebrates in
     three Michigan streams.   Tech. Rept. #20.  IWR.  Mich. St. Univ.

Go!ding, B. L.  1978.  Discussion of Proc.  Paper 13068.  ASCE J. Hydraul.
     Div.  103:458-459.

Graham, P. H., C.  S. Lawrence, and H. J. Mai Ion.  1974.  Estimation of
     imperviousness and specific curb length for forecasting stormwater
     quality and quantity.  Journal WPCF 46:717-725.

Grahn, 0.   1970.   Macrophyte succession in Swedish lakes caused by deposi-
     tion  of airborne acid substances.  Water, Air, and Soil Poll.  7:295-

Grenney, W. J., D.  B. Porcella, and M. L. Cleave.  1976.  Water quality
     relationships  to flow-streams and estuaries.  In:  Methodologies for
     the Determination of Stream Resource Flow Requirements:  An Assessment.
     USFWS.  Ft. Collins, CO 80526.  pp.  35-88.

Gundlach, D. L.  1976.  Unit hydrograph parameters versus urbanization.
     ASCE J. Irrig.  Drain.  Div. 102:388-392.

Guy, H. P., and D.  E. Jones, Jr.  1972.  Urban sedimentation -- In perspec-
     tive.  ASCE J.  Hydraul. Div.  98:2099-2116.

Haan, C. T., and R.  W. DeVore.   1978.   Proceedings of International Symposium
     on Urban Storm Water Management.   Univ. of Kentucky, Lexington, KY.
     348 p.

Haedrich, R. L.  1975.  Diversity and overlap as measure of environmental
     quality.  Wat.  Res.  9:945-952.

Haith, D. A.  1976.   Land use and water quality in New York rivers.  ASCE
     J. Env. Eng.  Div.  102:1-15.

Hamilton, M. A.  1975.  Indexes of diversity and redundancy.  Journal WPCF

Harrell, H. L.  1978.  Response of the Devil's River (Texas) fish community
     to flooding.   Copia.  1:60-68.

Hawkins, R. H.  1976.  Salt storage and runoff in urban watershed.  AS.CE
     J. Env. Eng.  Div.  102:737-743.

Hayden, J. W., J.  Dickson, and D.  Wisnuwski.  1972.  Water resources
     management for a small watershed in an urbanizing area.  In:  Watersheds
     in Transition.   A. C. Csallany et al. , eds.  Amer. Wat. Res. Assoc.
     Urbana, IL.  pp. 380-386.

Heck, K. L., Jr.  1976.  Community structure and the effects of pollution
     in sea-grass meadows and adjacent habitats.  Marine Biol.  35:345-357.

Hendrix, P.  1979.   Review of biological and ecological methods for assess-
     ing stream ecosystems.  Draft Report.   Urriv. of Georgia, Inst. of  Ecol.
     Athens, GA.

Hickock, E. A., M.  C. Hannaman, and N. C. Wenk.  1977.  Urban runoff treat-
     ment methods.   Volume I - Non-structural wetland treatment.  EPA-600/
     2-77-217.  U.  S. Environmental Protection Agency.  Cincinnati, OH.
     121 p.

Hocutt, C.  H.   1975.  Assessment of a stressed macro-invertebrate community.
     Wat. Res. Bull.  11:820-821.

Moiling, C. S.  1973.  Resilience  and stability of ecological systems.  Ann.
     Rev. Syst. & Ecol. 4:1-23.


Hossain, A., A.  R. Rao, and J.  W.  Delleur.   1978.  Estimation of direct
     ru'noff from urban watersheds.   ASCE J.  Hydraul.  Div.  104:169-188.

Howmiller, R. P., and M. A. Scott.   1977.  An environmental index based on
     relative abundance of oligochaete species.  Journal WPCF 49:809-815.

Hulbert, S. H.  1971.  The non-concept of species diversity:  A critique and
     alternative parameters.  Eool.   52:577-586.

International Hydrological Decade.   1974.  Hydrological effects of urbaniza-
     tion.  Report of the Sub-group on the Effects of Urbanization on the
     Hydrological Environment,  of the Co-ordinating Council of the Inter-
     national Hydrological Decade.   The Unesco Press, Paris, France.  280 p.

Israelsen, E. K., D. R. Bernard, R.  M. Twedt, and H.  M. Runke.  1975.  A
     technique for predicting the aquatic ecosystem response to weather
     modification.  PRWG138-1,  Utah Water Res. Lab.,  Utah State Univ., Logan,
     UT  84322.   157 p.

Jewell, T. K., T. J. Nunno, and D.  D. Adrian.  1978.   Methodology for cali-
     brating stormwater models.  ASCE J.  Env. Eng. Div. 104:485-501.

Jodie, J. B.  1975.  Quality of urban freeway storm water.  Trans. Res.
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Judd, J. H.  1970.  Lake stratification caused by runoff from street deicing.
     Water Res.   4:521-532.

Judd, J, and G.  A. Carlson.  1978.   Land use and water quality.  Clearwaters

Kaesler, R. L.,  E. E. Herricks, and J. S. Grossman.  1978.  Use of indices
     of diversity and hierarchical  diversity in stream surveys.  In:
     Biological Data in Water Pollution Assessment:  Quantitative and
     Statistical Analyses.  ASTM STP 652.  K. L. Dickson, J. Cairns, Jr.,
     and R. J. Livingston, eds.  pp. 92-112.

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     small watersheds.  [Quoted from Wanielista 1978].  USDA, SCS.  TP-149.

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Krishnan, K., P. Radha, J. J. Lizcano, L. E. Erickson, and L. T. Fan.  1974.
     Evaluation of methods for estimating stream water quality parameters
     in a transient model for stochastic data.  Wat.  Res.  Bull. 10:899-913.

Lager, J. A., R. P. Shubinski,  and L. W.  Russell.  1971.  Development of a
     simulation model for stormwater management.  Journal WPCF 43:2424-2435.


Lindholm, 0. G.  1976.   Pollutional  analysis of combined sewer systems.
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     McGraw-Hill  Book Co., San Francisco, CA  654 p.

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     sources.  Journal  WPCF 46:1849-1872.

Logan, D. T.  1977.  Discussion:   Indexes associated  with information theory
     in water quality.   S. M.  Zand,  Journal  WPCF 48:2026 [1976].   Journal
     WPCF 49:879-880.

Lowe, R. L.  1974.  Environmental requirements and pollution tolerance of
     freshwater diatoms.   EPA-670/4-74-005.   U. S. EPA, Cincinnati, OH 334 p.

Lynch, M. P.  1975.  The use of physiological indicators of stress in marine
     invertebrates as a tool for marine pollution monitoring.  In:  Pro-
     ceedings of Marine Technology Society,  Tenth Annual Conference.   NOAA
     75051202.  NTIS Springfield, VA.   22161.  pp. 881-890.

MacKenthun, J. L.   1975.   Legislative requirements.  In:  The Integrity of
     Water.  U.S.  EPA.   Sup. of Docu.   053-001-01068-1, Wash. DC  20402.
     pp. 5-8.

Mackenzie, M. J., and J.  V.  Hunter.   1979.   Sources and fates of aromatic
     compounds in urban stormwater runoff.   Env. Sci. and Tech.  13:179-183.

MacMahon, J. A.,  D. L.  Phillips,  J.  V.  Robinson, and  D. J. Schimpf.  1978.
     Levels of biological  organization:  An organism-centered approach.
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MacMahon, J. A.,  D. J.  Schimpf, D. C.  Anderson, K. G. Smith, and R. L. Bayn,
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     system concepts.  Unpublished manuscript.  Utah  State Univ., Logan,  UT
     84322.  30 p.

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     particular emphasis on recently colonized bodies of water.   In:  The
     Structure and Function of Freshwater Microbial Communities.   J. Cairns,
     Jr., ed.  Amer. Micros. Sco. Symp.  pp. 121-149.

Malone,  R.  F., D. S. Bowles, W. J. Grenney, and M. P. Windham.  1979.  Sto-
     chastic analysis for water quality.   Utah Water Res. Lab.  UWRL/Q-79/01,
     Utah State Univ.,  Logan, UT  84322.   75 p.

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Margalef, R.   1975.   External  factors and ecosystem stability.  Schweiz.  Z.
     Eydrol.  37:102-117.

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     Univ. Press,  Princeton, NJ.  235 p.

McElroy, A. D.,  S. Y. Chiu, J. W. Nebgen, A. Aleti, and F. W. Bennett.  1976.
     Loading functions for assessment of water pollution from nonpoint
     sources.   EPA-600/2-76-151.  U.S. EPA, Wash. DC  20460.  445 p.

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     J. Environ.  Qual.  1:86-89.

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     PRWG136-1,  Utah State Univ., Logan,  UT  84322.  228 p.

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     structure of  a woodland springbrook  community.  Ecol. 48:139-149.

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     land springbrook.  Hydrobiologia 32:305-339.

Mischi, E. F., and V. V.  Dhasmadhikari.  1971.  Runoff--A potential  resource.
     Water and Wastes Eng. 8:28-31.

Mitchell, D.,  and  J. C. Buzzell.  1971.  Estimating eutrophic potential of
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                                SECTION  IX


American Public Works Association.   1969.   Water pollution aspects of urban
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AWRA Research Committee.   1978.   Some research and information needs  in
     surface water quality.   Wat. Res.  Bull.  14:517-523.

Bailey, B. H., R. H.  Ramsey, B.  J.  Claborn, R. M. Sweazy,  and D.  M. Wells.
     1975.  Variation of urban runoff quality and quantity with duration and
     intensity of storms —  phase III:  final report;  Vol.  3.   OWRT Project
     No. B-177-TEX.  Water Resources Center; Texas Tech Univ., Lubbock, TX.
     24 p.

Bayne, B. L.  1973.  Aspects of physiological stress  in estuarine conditions.
     Biol. J. Linnean Soo.  6:360.

Beeton, A. M., P. K.  Kovacic, and A. S.  Brooks.  1976.   Effects of chlorine
     and sulfate reduction on Lake  Michigan invertebrates.   Center for Great
     Lakes Studies, University of Wisconsin-Milwaukee.   EPA 600/3-76-036.
     U. S. EPA Duluth, MN  55804.   132 p.

Bell, D. E., B. J. Claborn,  R. M.  Sweazy,  R. E. Peterson,  R.  H. Ramsey, and
     D. M. Wells.  1975.   Variation of urban runoff quality and quantity with
     duration and intensity of storms — phase III: final  report; Vol.  2.
     OWRT Project No. B-177-TEX. Water Resources Center;  Texas Tech  Univ.
     Lubbock, TX.  50 p.

Bergersen, E. P., and D.  L.  Galat.   1975.   Coniferous tree bark:   a light
     weight substitute for limestone rock  in barbeque basket microinverte-
     brate samplers.   Water Res. 9:729-731.

Betson, R. P., and W. McMaster.   1975.   Nonpoint source mineral water quality
     model.  Journal WPCF 47:2461-2473.

Birge, W. J., J. A. Black, A. G. Westerman, P. C. Francis  and J.  E.  Hudson.
     1977.  Embryopathic effects of waterborne and sediment accumulated
     cadmium,, mercury and zinc on reproduction and survival of fish and
     amphibian populations in Kentucky.   Water Resources Research Institute.
     University of Kentucky, Lexington,  KY.  28 p.

Birge, W. J., J. J. Just, A. G.  Westerman, J. A. Black, and 0. W. Roberts.
     1975.  Sensitivity of vertebrate embryos to heavy metals as a criterion


     of water1 quality.   Phase II.   Bioassay procedures using developmental
     stages as test organisms.   Water Resources Institute.  University of
     Kentucky.  Lexington, KY.   36 p.

Bloomfield, J. A., R. A. Park,  D.  Scavia, and C. S. Zahorcak.  1973.  Aquatic
     modeling in the eastern deciduous forest biome, U. S. International
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     pp. 139-158.

Bowdre, J. H., and N. R. Krieg.  1974.  Water quality rnonotoring:  bacteria
     as indicators.  Bulletin 69.   Virginia Water Resources Research Center,
     Virginia Polytechnic Institute and State University.   Blacksburg,
     VA.  20 p.

Brockway, C. E., and K. P. Grover.  1978.  Evaluation of urbanisation and
     changes in land use on the water resources of mountain valleys.  Idaho
     Water Resources Research Institute, Univ. of Idaho, Moscow, ID.  105 p.

Brownlee, R. C., T. A.  Austin,  and D. M. Wells.  1970.  Variation of urban
     runoff with duration and intensity of storms.   Interim report.   Water
     Resources Center;  Texas Tech Univ.  Lubbock, TX.  67 p.

Bucci, S. A., C. A. Carlozzi, and B. G. Oakland.  1976.  A modeling approach
     for selecting and measuring the ecological impacts of natural and
     cultural nutrient influences on lakes and their watersheds.  OWRT
     Pro. A-071-MASS.  Water Resources Research Center.  Univ. of Massachu-
     setts.  Amherst, MA  01002.  32 p.

Buikema,  A.  L.,  Jr., C. L. See, and J. Cairns, Jr.  1977.  Rotifer sensiti-
     vity to combinations of inorganic water pollutants.  Bulletin 92.
     Water Resources Research Center, Virginia Polytechnic Inst. and State
     Univ., Blacksburg, VA.  42 p.

Buikema, A. L. Jr., J.  Cairns,  Jr., and G. W. Sullivan.  (Undated)  Rotifers
     as monitors of heavy metal pollution in water.  Bulletin 71.  Water
     Resources Research Center, Virginia Polytechnic Inst. and State Univ.,
     Blacksburg, VA.  74 p.

Bureau of Water Management.  1971.  Chlorinated municipal waste toxicities
     to rainbow trout and fathead minnows.  EPA Grant No. 18050 Gil.  Michi-
     gan Dept. of Natural Resources.  Lansing, MI  48926.  49 p.

Cairns, J., Jr.  ed.  1971.  The structure and function of freshwater
     microbial communities.  Amer. Micros., ASTM, Philadelphia, PA.  301 p.

Cairns, J., Jr.  1974.   Environmental quality indicators for aquatic eco-
     systems.  In:  Environ. Plan. Sem.3 Proc.3 Drexel Univ., Philadelphia,
     PA.  pp. 47-61.

Cairns, J., Jr.  1974.  Indicator species vs. the concept of community
     structure as an index of pollution.  Wat. Res. Bull. 10:338-347.


Cairns, J., J. S. Crossman, K. L. Dickson, and E. E. Herricks.  1971.  The
     effect of major industrial spills upon stream organisms. 26th Industrial
     Waste Conference.   Purdue Univ.  Part One.  pp. 156-170.

Cairns, J., Jr., and K. L. Dickson.  1972.  An ecosystematic study of the
     South River, Virginia.  Bulletin 54.  Water Resources Research Center.
     Virginia Polytechnic  Inst. and State Univ., Blacksburg, VA.  104 p.

Cairns, J., Jr., K. L.  Dickson, and E. E. Herricks.  eds.  1975.  Recovery
     and restoration of damaged ecosystems.  University Press of Virginia.
     Charlottesville, VA.  531 p.

Cairns, J., Jr., G. R.  Lanza, and B. C. Parker.  1972.  Pollution related
     structural and functional changes in aquatic communities with emphasis
     on freshwater algae and protozoa.  Proc.  Acad. Nat. Sci. Phila. 124:

Canale, R. P.  1976.  Modeling biochemical processes in aquatic ecosystems.
     Ann Arbor Science Pub!., Ann Arbor, MI.  390 p.

Carey, G. W., L. Zobler, M. R. Greenberg, and R. M. Hordon.  1972.  Urbaniza-
     tion, water pollution, and public policy.  Center for Urban Policy
     Research, Rutgers, Univ., The State Univ. of New Jersey, New Brunswick,
     NJ.  214 p.

Carlsson, L., and J. Falk.  1978.  Urban Hydrology in Sweden - an inventory
     of the problems and their costs.  Report no. 3017.  Department of Water
     Resources Engineering, Lund Institute of Technology, University of Lund.

Chang, S. L.  1972.  Zoomicrobial indicators of water pollution.  National
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                                  TECHNICAL REPORT DATA
                           (Please read Instructions on the ft'icisc before completing)
                                          6. PERFORMING ORGANIZATION CODE
      Donald  B.
      Darwin  L.
      Utah Water  Research Laboratory
      Utah State  University
      Logan Utah   84322
                                                           8. PERFORMING ORGANIZATION REPORT NO.
                                                           10. PROGRAM ELEMENT NO.
      Environmental  Research Laboratory--Corvallis
      Office of  Research  and Development
      Environmental  Protection Agency
      Corvallis  Oregon  97330
                                                           T. RECIPIENT'S ACCESSION NO.
                                                           5. REPORT DATE
                                          1 1. CONTRACT/GRANT NO.
                                                             Req.  CC81224-J
                                          13 TYPE OF REPORT AND PERIOD COVERED
                                          _ Final - Literature  review
                                          14. SPONSORING AGENCY CODF

      Project Officer:  Kenneth W.  Malueg, Environmental  Research Laboratory
                        Corvallis.  Oregon 9733Q   (503)757-4761  (FTS) 420-4761
 Literature on urban  nonpoint source runoff was surveyed  to  determine the magnitude of
 the effects of  that  source of contaminants to stream  ecosystems.   Very little informa-
 tion was available on  ecosystem effects although extensive  literature on all aspects
 associated with contaminant loading was found.  However,  urban NPS runoff probably
 exerts unique effects  on  stream communities because of its  random magnitude/impulse
 loading of contaminants.   Because control of NPS runoff  is  expensive, it is important
 to determine its  actual  impacts on stream communities.   Ecological literature provided
 a basis for evaluating such impacts based on benthic  invertebrate biomass and diversity
 measurement of  community  primary production and respiration,  carbon cycling, and vari-
 ables related to  the contaminant concentrations in the stream.  We concluded that a
 stochastic approach  for assessing the impacts would be most feasible for evaluating
 impacts of urban  NPS runoff.
                               KEY WORDS AND DOCUMENT ANALYSIS
 Water qua!i ty
 Stream ecosystems

 Release to public
                                              b IDENTIFIERS/OPEN ENDED TERMS
                               Urban nonpoint source
                             19. SECURITY CLASS (This Report)

                             _ Unclassified,
                             20 SECURITY CLASS (This page)
c. COSATI Field/Group

                                                                         21. NO. OF PAGES
                                                                         22. PRICE
EPA Form 2220-1 (Rev. 4-77)