JBRM2012

This article was downloaded by: [Universiti Tunku Abdul Rahman], [ChewKhun Tan]
On: 16 December 2012, At: 19:13
Publisher: Taylor & Francis
Informa Ltd Registered in England and Wales Registered Number: 1072954 Registered office: Mortimer
House, 37-41 Mortimer Street, London W1T 3JH, UK
International Journal of River Basin Management
Publication details, including instructions for authors and subscription information:
http://www.tandfonline.com/loi/trbm20
Urban runoff impacts on receiving aquatic
ecosystems assessed using periphyton community
Tan Chew Khun
a
, Carolyn Oldham
b
& Louis Evans
c
a
Faculty of Engineering and Green Technology, Universiti Tunku Abdul Rahman, Bandar
Barat, Kampar, 31900, Perak, Malaysia
b
School of Environmental Systems Engineering, University of Western Australia, Western
Australia, Australia E-mail:
c
Curtin University of Technology, Western Australia, Australia
Accepted author version posted online: 10 Apr 2012.Version of record first published: 03
May 2012.
To cite this article: Tan Chew Khun, Carolyn Oldham & Louis Evans (2012): Urban runoff impacts on receiving aquatic
ecosystems assessed using periphyton community, International Journal of River Basin Management, 10:2, 189-196
To link to this article: http://dx.doi.org/10.1080/15715124.2012.683007
PLEASE SCROLL DOWN FOR ARTICLE
Full terms and conditions of use: http://www.tandfonline.com/page/terms-and-conditions
This article may be used for research, teaching, and private study purposes. Any substantial or systematic
reproduction, redistribution, reselling, loan, sub-licensing, systematic supply, or distribution in any form to
anyone is expressly forbidden.
The publisher does not give any warranty express or implied or make any representation that the contents
will be complete or accurate or up to date. The accuracy of any instructions, formulae, and drug doses
should be independently verified with primary sources. The publisher shall not be liable for any loss, actions,
claims, proceedings, demand, or costs or damages whatsoever or howsoever caused arising directly or
indirectly in connection with or arising out of the use of this material.

Research paper
Urban runoff impacts on receiving aquatic ecosystems assessed using periphyton
community
TAN CHEW KHUN, Faculty of Engineering and Green Technology, Universiti Tunku Abdul Rahman, Bandar Barat,
Kampar 31900, Perak, Malaysia. Email: tckhun@utar.edu.my (Author for correspondence)
CAROLYN OLDHAM, School of Environmental Systems Engineering, University of Western Australia, Western
Australia, Australia. Email: carolyn.oldham@uwa.edu.au
LOUIS EVANS, Curtin University of Technology, Western Australia, Australia
ABSTRACT
Urban runoff is a major cause of stream degradation. For appropriate management action, there is a need to establish whether or not the runoff is degrad-
ing or potentially will degrade the receiving water. This work explored the use of local periphyton communities for assessing the ecological impacts of
runoff of two urban catchments. Periphyton communities were colonized on glass substrate for 2 weeks in riverine waters receiving the urban runoff and
in non-receiving waters at an upstream reference site. The receiving communities were evaluated against the reference community for photosynthetic
efficiency and tolerance to copper, identified to be a significant runoff stressor. Photosynthesis efficiency was measured as a PSII quantum yield, and
community tolerance was assessed using a laboratory ecotoxicological test. The hypothesis tested is that the runoff degrades by causing communities in
receiving waters to become more sensitive to runoff stressors. The bioassessment indicated a much higher copper threshold than provided for in the
generic water quality guidelines. The significance in the use of local community is that it allows for ecological risk analysis of exposure to runoff stres-
sors, providing site-specific information relevant to management.
Keywords: Community tolerance; runoff; ecological risk; periphyton; photosynthesis; efficiency
1 Introduction
Globally, urban streams are being contaminated and degraded by
catchment runoff (Ellis and Hvitved-Jacobsen 1996). Before
implementing any management strategies, it needs to be estab-
lished that the runoff indeed is or potentially is degrading to the
receiving aquatic ecosystems. Conventionally, the assessment is
by comparing runoff physico-chemical water quality against
nationalguidelines.However,thisassessmentmethodhasbeencri-
ticized as being unrealistic. First, guideline criteria are generic,
hence may not be relevant to the site of concern. The literature
data used for deriving the guideline criteria are based on laboratory
ecotoxicological tests that were not designed for local biotic or
abiotic conditions. Second, there is concern in the way results of
single species and lower levels of biological organization are
extrapolated for use for multiple species at higher levels (United
States Environmental Protection Agency 1986, OECD 1992).
Extrapolation using arbitrary safety factors lacks scientific basis,
and there has been little work done to evaluate the effectiveness
of such an approach in providing a desired level of protection
(Forbes and Forbes 1994). Furthermore, field effects may not be
directly related to toxicity but more to conditions of exposure
and physiology (Ellis et al. 1995). Environmental factors are
known to affect exposure by modifying the chemical forms or
uptake processes of contaminants (Ratte 1999, DiToro et al.
2000). Standard toxicological testing makes use of specified test
biota. However, very often the specified test biota do not resemble
the local biota of concerned in make-up, physiology, or exposure-
conditioning (Suter II et al. 1985). This can result in the setting of
the safety level that is not effective in protecting the local species.
Exposure conditions to runoff are complex and highly variable in
composition and magnitude of stressors (Ellis et al. 1995). While
biological organisms are likely to adapt and acclimatize to long-
term gradual exposure, they may not readily adapt to the short-
term highly variable exposure of runoff (Horner 1995). Addition-
ally, a mixture of two or more stressors may be synergistic or
Received 20 January 2012. Accepted 3 April 2012.
ISSN 1571-5124 print/ISSN 1814-2060 online
http://dx.doi.org/10.1080/15715124.2012.683007
http://www.tandfonline.com
189
Intl. J. River Basin Management Vol. 10, No. 2 (June 2012), pp. 189–196
# 2012 International Association for Hydro-Environment Engineering and Research
Downloadedby[UniversitiTunkuAbdulRahman],[ChewKhunTan]at19:1316December2012

antagonistic (Blanck et al. 1988). For management to be effective,
the link between stressor(s) and effect(s) needs to be identified and
understood. It is not practical to evaluate the toxicity for each and
every chemical of a mixture on each and every biological species,
let alone the factorial of variation in chemical composition. Model-
ling approaches have been particularly useful for complex systems
and several models, such as the Biotic Ligand Model (DiToro et al.
2000) have been developed for analysis of the exposure com-
ponent. However, most of these models were developed for unin-
terrupted exposure conditions, and may not be readily applied to
the short-term, variable exposure conditions of runoff.
Another common assessment method is in situ biomonitoring.
It provides information about the spatial and temporal distri-
bution of local biological species. Although the biotic data can
be correlated to the physico-chemical data of the receiving
waters, they cannot be used to establish causality (United States
Environmental Protection Agency 2000), hence making it diffi-
cult to target management actions. Moreover, the observed
biotic variation is essentially associated with quasi steady-state
conditions, and may not reflect the short-term effects of urban
runoff (Yasuno and Whitton 1988). Interpreting biomonitoring
results requires caution since not all changes in community struc-
ture are due to anthropogenic degradation, especially where
runoff is seasonal. Natural seasonal changes can often bring
about similar changes (Gray 1980). A more realistic and effective
way to assess degrading effects of urban runoff is by measuring
the response indicator of a biological community local to the
site. The risk of degradation will be high if the urban runoff
affects the ability of the community to maintain its ecological
function. The periphyton community is an important ecological
component of aquatic ecosystems, with an essential function of
primary production by photosynthesis. The hypothesis is that
urban runoff degrades the receiving aquatic ecosystem by
causing the periphyton community to change to one that is
more sensitive to the runoff stressor(s) and also less efficient in
photosynthesis. The hypothesis was tested using two adjacent
urban catchments of the Bayswater (BW) main drain (MD) and
Chapman (CM) MD in Western Australia and an upstream refer-
ence site on the adjoining reach of the estuarine Swan River. The
significance in the use of local biological community is that it pro-
vides site-specific information and allows for ecological risk
analysis of exposure to specific runoff stressors, all of which
are relevant to effective management.
2 Materials and methods
2.1 Study site
The catchments of the BWMD and CMMD are located adjacent
to each other, north-east of metropolitan Perth, Western Australia
(Figure 1). The BWMD catchment is almost 10 times larger in
area than the CMMD catchment (Table 1). Both catchments
are extensively urbanized, with mixed landuse of residential,
industrial, commercial and recreational. The drainage system
was constructed for conveyance of storm runoff from the catch-
ments for discharge into the Swan River estuary.
The climate is Mediterranean: wet, cold in winter and dry,
warm in summer. The catchment soils are predominantly Bas-
sendean sand (Davidson 1995), which is highly porous, low in
retention capacity for metals, and has little acid buffering
capacity (Newman and Marks 1979). The groundwater is gener-
ally at less than 30 m above the sea level and the flow is towards
the Swan River estuary (Hirschberg and Appleyard 1996). Both
the BWMD and CMMD have been rated among the top five con-
tributors to contamination of the Swan River estuary (Henderson
and Jarvis 1995). Contamination sources in the catchments have
been identified and are related to current and past landuses. Metal
sources are mainly in the industrial area (Klemm and Deeley
1991, McCarthy and Nicolson 1996). In addition to current
industrial activities, a major contributor of metal contamination
includes several industrial waste landfills, notably a stockpile
of pyritic material occupying 0.17 km2
area of the CMMD catch-
ment (ERS 1998). The metals found in the pyritic material
include arsenic, cadmium, copper, lead, nickel and zinc.
Figure 1 A sketch map of the catchments of the BWMD and CMMD
showing the locations of the monitoring stations (S1 and S2) and periph-
yton colonization (S3–S5)
Table 1 Catchment characteristics of the BWMD and CMMD
BW CM
Whole catchment area (km2
) 27.2 2.95
Catchment area monitored (km2
) 26.3 2.95
Impervious area (% of catchment area) 97 56
Population density (km2
) 1729 1369
Landuse (% of catchment area)
Residential 79 56
Industrial 11 28
Commercial 2 5
Parkland and reserve 8 11
Road and railway (km) 409 107
Mean annual temperature (8C) 18.1 18.1
Mean max/min temperature (8C) 30.6/8.3 31/7.9
Annual precipitation (mm) 718 709
190 Tan Chew Khun et al.

2.2 Water quality monitoring
Water quality of the MDs was monitored about 1 km upstream of
the drain confluences with the receiving Swan River estuary (S1
and S2 in Figure 1). Not included in the monitoring was the lower
subcatchment of the BWMD since it joins the MD after the moni-
toring point, i.e. 97% of the catchment flow was monitored. The
monitoring system comprised a controller unit (Campbell Scien-
tific CR10X data logger), an assembly of four sensors for in situ
measurement of the water level, water temperature cum electrical
conductivity (EC) (Campbell Scientific 247-L Probe), turbidity
(Analite 195/1/30), and pH (Greenspan Model PH100), and an
autosampler (ISCO 3700 Standard Model) for sampling water
for metal contaminants. Measurements of in situ variables were
logged as running averages of 1 min on the rising limb of the
hydrograph, 5 min on the falling limb, and 15 min for all other
conditions. Runoff discharge was determined from a calibrated
water level–discharge rating curve of the respective drains. Cali-
bration data of water level and runoff discharge for CMMD were
determined at a conduit just downstream of the monitoring site,
using a flow current meter (Montec DETEC 3013). The cali-
bration data for BWMD were acquired from the Water and
Rivers Commission of Western Australia, who has a monitoring
station at the study site. It employs a weir system for measuring
the discharge. Water sampling for metals was triggered at the
start of storm events; thereafter at every 3–5 cm change in
water level. A total of 17 water samples were collected for
BWMD and 7 samples for CMMD. They were analysed for
Al, As, Cd, Cr, Cu, Mn, Ni, Pb and Zn by inductively coupled
plasma mass spectrometry after digestion in nitric acid (HNO3
AJAX Analytical Reagent). The loading rate of metal was esti-
mated as the product of concentration and water discharge rate.
Water quality was monitored as part of a 2-year programme
from May 2000 to December 2001. For the purpose of this
study, the data covering the periphyton colonization period
from 21 September to 5 October 2001 was taken for subsequent
analysis.
2.3 Field colonization of periphyton communities
Periphyton communities were colonized on the Swan River at
three sites, about 2.5 km apart (Figure 2). The drain sites S3
and S4 were located about 5 m downstream of the drain con-
fluences with the river. The reference site S5 was located
upstream from the drain sites. With no other MDs along the
river reach between the sites, the difference in riverine conditions
between drain sites and the reference site was assumed to be due
to the respective drain discharge.
At each site, three lots of five collectors were placed 3 m apart
across the river flow and 5 m from the northern bank, each lot
being held suspended at 1 m depth by a float and moored in pos-
ition by weights (Figure 2). Each lot represented a replicate
sampling. Each collector comprised 10 glass slides (75 mm ×
25 mm × 1 mm), held in a plastic rack. After colonization for
2 weeks (21 September to 5 October 2001), the developed per-
iphyton communities were collected for laboratory analysis.
2.4 Assessment against generic water quality guidelines
The physico-chemical variables of the drain discharge were
assessed against the national water quality guidelines
(ANZECC 2000) for lowland river systems of the southwest
region of Australia. As the catchments of the BWMD and
CMMD are extensively developed, they were regarded as
highly disturbed ecosystems, hence the default trigger values
for 80% species protection were applied. Assessment for poten-
tial stressors was based on the quotients of the median values of
the water quality variables to the guideline limits (Sample et al.
1998). For pH, where a lower value is more adverse, the inverse
quotient value was taken. Variables with quotients .1 (in the
case of pH, the inverse quotient .1) were regarded as stressors.
The stressor with the highest quotient would be the most sig-
nificant. In a mixture of stressors, the resultant effect can be
antagonistic or synergistic; synergistic effect can be additive or
multiplicative (Blanck et al. 1988, Fukunaga 2010). For simpli-
city, an additive effect is assumed (ANZECC 2000, Kamo and
Nagai 2008). The sum of stressor quotients was taken as an indi-
cator of the runoff stress level.
2.5 Assessment using local periphyton communities
The periphyton communities were assessed for differences in
biomass, photosynthesis efficiency, and tolerance to the most sig-
nificant runoff stressor. The community biomass was determined
as ash-free dry weight according to the APHA (1998) Method
10200. Photosynthesis efficiency was determined as the PSII
quantum yield by the pulse amplitude modulation fluorometry
(WALZ Underwater Diving PAM). The periphyton were
removed from the glass substrate, homogenized and maintained
in 100 ml of culture medium. The culture medium was f/2
(Guillard 1975), modified by reducing the copper and zinc con-
centration to 10% of the original recipe, to a final concentration
of 0.34 mg Cu/m3
and 0.77 mg Zn/m3
, and without the ingredient
Figure 2 Schematic assembly of glass substrate for periphyton coloni-
zation in the river
Urban runoff impacts assessed using periphyton community 191

ethylenediaminetetraacetate. The diluent water was the filtered
(0.45 mm) natural seawater (Indian Ocean, Western Australia),
diluted with deionized water to 10‰ salinity (the mean salinity
of the estuarine water, http://www.wrc.wa.gov.au/srt/
riverscience/physical). Aliquots of 0.5 ml of test periphyton
samples, in triplicate, were diluted with 4 ml of the culture
medium in culture plates (5-ml × 24-well, clear polystyrene,
straight walls, flat base). Photosystem II yield was determined
after conditioning at 208C for 24 h of 12 h dark: 12 h light
(60 W OSRAM 3000 lux daylight fluorescent light tube). The
dilution was necessary so as to obtain a thin dispersion of the
organisms fully covering the cell floor for the measurement.
Community tolerance to copper was determined using a short-
term ecotoxicological test. The periphyton samples were
exposed, in triplicate, for 24 h to a series of copper concen-
trations of 0.4, 2, 4, 10, 50, 100 and 250 mg Cu/l (Analytical
Reagent CuSO4.5H2O, in the deionized water). The ratio of the
sample to the copper solution was 0.5:4 ml, the same as for com-
munity PSII measurement above. The test samples were simi-
larly conditioned as for the PSII measurement above. A control
was included. Photosystem II yield was determined at the end
of the exposure. The effect concentrations at inhibition of 1%
(EC1), 5% (EC5) and 50% (EC50) were determined by the
ICp approach (Guillard 1975). These effect endpoints corre-
spond to the protection of the community photosynthesis func-
tioning at the level 95% (EC5), 99% (EC1) and 50% (EC50)
respectively. The threshold concentration or no-effect concen-
tration was taken to be EC5. The community tolerance was
also characterized by the parameter of tolerance width, defined
by the equation:
Tolerance width (mg Cu/l) ¼ EC50–EC5 (1)
The tolerance width provides an indication of the copper concen-
tration above the threshold that the community can tolerate
before its photosynthesis performance declines by 50%. Here,
the background concentration need not be known a priori, and
the management needs only focus on the concentration incre-
ment. Between community differences were evaluated by non-
parametric tests of significance such as the median test,
Kruskal–Wallis and Mann–Whitney.
3 Results
During the field colonization from 21 September to 5 October
2001, there were two series of storm events separated by at
least 4 days of non-events. The BWMD and CMMD displayed
similar water flow pattern, differing only in magnitude. In
general, the flow in the BWMD was higher by an order of mag-
nitude (Figure 3).
3.1 Assessment by generic water quality guidelines
The discharge of both catchments displayed characteristics
typical of the urban runoff, i.e. highly variable and multiple stres-
sors. The stressors with median values exceeding the guidelines
for both catchments were of the same types, namely EC, turbid-
ity, Al, Cu and Zn (Table 2). These stressors were higher in mag-
nitude in the CMMD than in the BWMD, except for Al which
was equally high in both drains. The sum value of exceedance
quotients was higher for the CMMD than for the BWMD, by
more than a factor of 2.
As the quotient and CM:BW ranked the highest for copper
compared to the others (quotient of 47 and CM:BWof 9, respect-
ively, Table 2), copper was selected for the subsequent ecotoxi-
cological test.
3.2 Assessment using local periphyton communities
Biomass was comparable between the communities of the
BWMD and reference sites, but was lower for the community
of the CMMD site (Table 3). However, statistically the differ-
ences between all communities were not significant (Mann–
Whitney sig . 0.5). In contrast, photosynthesis efficiency was
significantly lower for the receiving communities than for the
reference non-receiving community (Mann–Whitney sig ,
0.001). Between the communities receiving runoff, there was
no significant difference (sig . 0.5). The ecotoxicological test
indicated increased deviation in response between the receiving
communities and the reference community with increased
copper concentration (Figure 4). Both the receiving communities
deviated to lower yields. The exposure concentration at which
the community declined in yield by 1% (EC1), 5% (EC5) or
50% (EC50) from the initial yield was in the order of CM ,
BW , reference, respectively (Table 3). The tolerance width
was also in the same order.
Figure 3 Hydrographs of the BWMD and CMMD covering the field
colonization period 21 September to 5 October 2001

4 Discussion
The periphyton communities colonized in waters receiving runoff
from the BWMD and CMMD were less efficient in photosynthesis
and more sensitive to copper toxicity than the reference
community. In-stream hydraulics have been cited as a control on
the biomass of the periphyton community (Weitzel 1979, Peterson
and Stevenson 1990, Rosen 1994). As no significant differences in
biomass were observed between the receiving communities and
the non-receiving reference community, hydraulic effects of the
drain runoff could not be controlling. This finding has provided
some validity to the assumption of similar background riverine
conditions at the drain sites and reference site.
The CMMD community was less tolerant or more sensitive to
copper than the BWMD community. With the reference commu-
nity accounting for background effects, the observed differences
between the drain sites would be attributed to the difference in
the drain runoff. Since the runoff volume from the CMMD
was significantly lower than from the BWMD, runoff quantity
would not be a contributing factor. That leaves runoff quality
as the likely controlling factor. The CMMD that was poorer in
water quality than the BWMD would be expected to have a
greater impact.
This study used only one reference site which may be limiting
especially when it was not nearby but at least 2.5 km away from
the drain sites. The experimental design can be improved by
including additional reference sites immediately upstream of
the drain outfalls to better define background conditions closer
Figure 4 Response curves to 24-h copper toxicity test of the periphy-
ton communities receiving runoff from the BWMD and CMMD, and
non-receiving at an upstream reference site of the Swan River estuary.
Response was the community photosynthesis efficiency as the mean
PSII quantum yield, error bars are +1 standard deviation (n ¼ 3)
Table 2 Comparison of the physico-chemical water qualities of the BWMD and CMMD against the ANZECC (2000) guidelines for freshwater
systems
Water quality parameter MD No. of cases Median Range CM/BW Quotient (median/guideline)
EC (mS/cm)
BW 1780 0.52 0.50–0.92
1.3
1.7
CM 1571 0.69 0.52–0.81 2.3
Turbidity (NTU)
BW 1780 30 11–1200
2.1
1.5
CM 1571 64 3.2–1300 3.2
Al (mg/l)
BW 17 3.1 1.3–7.8
0.3
21
CM 7 1.0 0.8–2.2 6.8
Cu (mg/l)
BW 17 0.013 0.008–0.025
9.0
5.3
CM 7 0.12 0.058–0.200 47
Zn (mg/l)
BW 17 0.11 0.08–0.16
3.4
3.6
CM 7 0.37 0.13–0.55 12
Sum of quotients . 1
BW
2.2
33
CM 71
Note: Listed are those variables whose median values exceeded the guidelines limits.
Table 3 Effect endpoints of biomass and photosynthesis yield of the field communities to copper toxicity
Reference BW CM
Total biomass (mg) as AFDW 14 + 2 14 + 2 10 + 4
Initial PSII yield (Y0) 0.690 + 0.002 0.669 + 0.008 0.670 + 0.003
EC1 (mg Cu/l) 3.4 + 1.5 1.9 + 1.4 0.8 + 0.4
EC5 (mg Cu/l) 8.5 + 1.2 4.7 + 0.6 1.9 + 0.1
EC50 (mg Cu/l) 62 + 3 33 + 3 9 + 3
Tolerance width (mg Cu/l) ¼ EC50–EC5 53 28 7
Note: Values are mean + 1 s, n ¼ 3; AFDW is ash-free dry weight; EC is effect concentration.

to the drain outfalls. Most researchers reported increased com-
munity tolerance to a baseline environmental exposure (Forbes
and Forbes1994 and references therein), which is supported by
the concept of pollutant-induced community tolerance (Blanck
et al. 1988). Only a few reported increased sensitivity (Crane
1990, Pennington and Scott 2001), concurring with this study.
The differences can be attributed to the different exposure con-
ditions of the studies. This study was under storm event con-
ditions, whereas most of the former studies were under less
variable conditions.
Increased sensitivity to a significant runoff stressor has impli-
cations for management. First, the ability to withstand further
stress would likely deteriorate for the sensitized community,
and a stressed community may not be fit to perform its usual
functions. If the conditions were prolonged or the community
did not recover quickly enough before the next storm event,
the community can enter a diseased state. As photosynthesis is
an important ecological function of primary productivity, a
decrease would certainly be degrading (Schindler 1996). The
risk of degradation by runoff from the CMMD was higher
since the receiving community was made more sensitive to the
significant stressor copper compared to the other communities.
As ecosystems tend to maintain functional status by means of
various compensatory mechanisms (Pratt and Cairns 1996, Sal-
minen et al. 2001), a significant decrease in function may be
indicative that the system is failing. As runoff stress tends to
be short-term, the ecosystem may recover when the stress is
removed during the dry season. However, in the case of the
study site, there have been reports of symptoms of degrading
ecological health such as increased frequency of algal blooms
and fish kills (WAWC 1994). This would be indicative of a
degrading ecosystem. Such a condition is known to be the case
with most urban streams; a condition aptly termed ‘urban syn-
drome’ (Cottingham 2004). Second, there is an implication con-
cerning the setting of protection limits. The usual approach for
setting the limit is to derive a concentration limit based on a
control or reference site. The concentration limits for copper
necessary for 99% and 95% protection of species were estimated
to be 3.4 and 8.5 mg Cu/l, respectively, for the reference site. For
waters receiving runoff from say the CMMD, the limits were 0.8
and 1.9 mg Cu/l, respectively, which were lower by an order of
magnitude than for the reference site. Another consideration is
the concentration width above threshold that the community
can tolerate. The tolerance width was narrower (by a factor of
2–9) for the drain sites than for the reference site. This difference
in protection limit has a significant implication for management.
If following the conventional approach, the use of the reference
limits would be under protective for the runoff impacted sites.
Both the assessment methods of using local communities and
using national guidelines rated the CMMD poorer in water
quality than the BWMD. The water quality monitoring was
able to characterize the trends and variability of the runoff quan-
tity and physico-chemical qualities. However, it could not be
used to establish the link to the observed biotic effects necessary
for effective management. On the other hand, the use of local per-
iphyton communities provided site-specific information data for
linking the observed biotic effects to the runoff. It also helped
quantify the risks of exposure to copper, identified as the most
significant runoff stressor by the guideline approach.
The underlying principle employed for the ecotoxicological
test is similar to that of the water quality guidelines. The differ-
ence between the two assessment methods would be related to
the specificity and relevance of the data to the study site. For
example, the 99% or 95% protection level for copper derived
using local community was much greater (by at least 103
times) than provided for by the guidelines. Taking the reference
site as an illustration, the values (3.4 and 8.5 mg Cu/l, respect-
ively) are three orders of magnitude higher than the guideline
limits (they are: 0.001 and 0.0014 mg Cu/l, respectively, for
freshwater systems, 0.0003 and 0.0013 mg Cu/l, respectively,
for marine systems; ANZECC 2000).
The implication is that the guideline limits are overly protec-
tive, at least for the study site. The generic guideline values are
derived using chronic test data and a safety factor of 2
(ANZECC 2000). Furthermore, the use of generic guidelines is
limited to allowing the probability estimation of exceedance fre-
quencies. It does not provide any mechanisms for assessing
impacts of multiple stressors. So, although both the study
drains had the same number and types of stressors, the stressors
differed in magnitude. Therefore, their impacts should be differ-
ent. The use of an exceedance quotient did provide a means of
weighing the impact factor. By assuming additive effects, the
sum quotients did provide an indication of the relative impact
factor. The impact from the CMMD was rated greater than
from the BWMD. This was corroborated by the ecotoxicological
assessment. In addition to providing site-specific information
and charting tolerance distribution, the use of local communities
can offer several other advantages. One advantage would be pro-
viding early warning to trigger management action. It has been
argued that the multispecies community would be more respon-
sive to stress than a single species could (Blanck et al. 1988).
From a management perspective, community-level response is
more relevant than response of individual organisms or species
level. Ecosystems are what management concerns with mainly,
and as communities are only one organizational level lower,
their response would provide an appropriate early warning to
the management. The current method of using community PSII
quantum yield has been demonstrated to be sensitive to short-
term stress of urban runoff. An intact natural community may
have provided the extra protection (Barranguet et al. 2000,
Dryden et al. 2004). Modifying factors at the study sites such
as high salinity and water hardness of the estuarine water are
known to have a sequestering effect on copper toxicity
(DiToro et al. 2000). This again highlights the need for site-
specific information.
Additionally, the collected communities can be used for other
analysis such as taxonomic analysis for community structure and
tissue residue analysis for evidence of exposure to particular

stressor(s). Evidence of exposure is important for identifying
stressor(s) and establishing causality (United States Environ-
mental Protection Agency 2000). The analysis of ecological
risks can be extended to other runoff stressors. The combined
information on structural and functional changes will provide a
better understanding of their relevance to runoff stressor(s).
Historically, it has been shown that an integration of as much
information as possible is necessary for development of effective
management strategies. The information that can be generated
with the use of local communities will certainly facilitate
better, more effective management of urban runoff.
5 Conclusion
The analysis of local periphyton community was demonstrated
to be an effective way for assessing runoff effects. It facilitated
ecological risk analysis of copper, identified as a significant
runoff stressor. The ecotoxicological data linked the physico-
chemical quality of the runoff to the observed changes in receiv-
ing communities effectively established the degrading effects of
the runoff. The information generated was site-specific and
therefore relevant to runoff management.
Acknowledgements
The authors acknowledge the financial support from the Bays-
water City Council of Western Australia, Australia.
References
ANZECC, 2000. The national water quality management strat-
egy paper no. 4: Australian and New Zealand guidelines for
fresh and marine water quality. vol. 1. Canberra: Australia
and New Zealand Environment Conservation Council
(ANZECC, A-21–A-22).
APHA, 1998. Standard methods for the examination of water
and wastewater. 20th ed. Washington, DC: American Public
Health Association (APHA), American Water Works Associ-
ation, and Water Environment Federation.
Barranguet, C., et al., 2000. Short-term response of monospe-
cific and natural biofilms to copper exposure. Journal of
the European Phycology, 35, 397–406.
Blanck, H., Wangberg, S.-A., and Molander, S., 1988. Pol-
lution-induced community tolerance: a new ecotoxicological
tool. In: J. Cairns, Jr and J.R. Pratt, eds. Functional testing
of aquatic biota for estimating hazards of chemicals. Phila-
delphia: American Society for Testing and Materials,
219–230.
Cottingham, P., 2004. World experience focuses on streams suf-
fering ‘urban syndrome’. Watershed, The Cooperative
Research Centre for Freshwater Ecology, October, pp. 4–5.
Crane, M., 1990. A review of the use of aquatic multispecies
systems for testing the effects of contaminants. Buckingham-
shire, UK: Department of Environment.
Davidson,W.A., 1995. Hydrogeology andgroundwater resources
of the Perth region, Western Australia. Perth: Geological
Survey of Western Australia.
DiToro, D.M., Allen, H.E., Bergman, H.L., Meyer, J.S., Santore,
R.C., and Paguin, P., 2000. The biotic ligand model: a compu-
tational approach for assessing the ecological effects of
copper and other metals in aquatic systems. New York: Inter-
national Copper Association Ltd.
Dryden, C.L., Gordon, A.S., and Donat, J.R., 2004. Interactive
regulation of dissolved copper toxicity by an estuarine
microbial community. Limnol Oceanogr, 49, 1115–1122.
Ellis, J.B. and Hvitved-Jacobsen, T., 1996. Urban drainage
impacts on receiving waters. Journal of Hydraulic Research,
34, 771–783.
Ellis, J.B., Shutes, R.B., and Revitt, D.M., 1995. Ecotoxicologi-
cal approaches and criteria for the assessment of urban runoff
impacts on receiving waters. In: E.E. Herricks, ed. Stormwater
Runoff and Receiving Systems. Boca Raton, FL: CRC Press,
113–125.
ERS, 1998. Centurion northwest: Tonkin Park stage II, Bassen-
dean – request for section 46 changes to minister’s conditions
and proponent commitments, June 1998. Perth: Environment
Risk Solution (ERS) Pty Ltd.
Forbes, V.E. and Forbes, T.L., 1994. Ecotoxicology. Theory and
in practice. London: Chapman & Hall.
Fukunaga, A., 2010. Effects of multiple heavy metals on estuar-
ine communities. Thesis (PhD). New Zealand: University of
Auckland.
Gray, J.S., 1980. The measurements of effects of pollutants on
benthic communities. Rapp. P.-v. Reun. Cons. Int. Explor.
Mer., 179, 188–193.
Guillard, R.R.L., 1975. Culture of phytoplankton for feeding
marine invertebrates. In: W.L. Smith and M.H. Chanley, eds.
Culture of marine invertebrate animals. New York: Plenum
Press, 29–60.
Henderson, R. and Jarvis, B., 1995. Urban drainage water pol-
lution assessment study. Report SSB 10/95. Water Authority,
Perth, Western Australia, Australia.
Hirschberg, K.-J.B. and Appleyard, S.J., 1996. A baseline survey of
non-point source groundwater contamination in the Perth Basin,
WesternAustralia.Perth:GeologicalSurveyofWesternAustralia.
Horner, R.R., 1995. Toward ecologically based urban runoff
management. In: E.E. Herricks, ed. Stormwater runoff and
receiving systems. Boca Raton, FL: CRC Press, 365–377.
Kamo, M. and Nagai, T., 2008. An application of the biotic
ligand model to predict the toxic effects of metal mixtures.
Environmental Toxicology and Chemistry, 27 (7), 1479–1487.
Klemm, V.V. and Deeley, D.M., 1991. Water quality in the Bays-
water main drain: a working paper for the Bayswater Inte-
grated Catchment Management Steering Committee. Report
No. 4 Nov 1991. Perth: Swan River Trust.

McCarthy, M. and Nicolson, J., 1996. Town of Bassendean:
results of 1995 drainage monitoring program. Western Aus-
tralia, Australia: Eastern Metropolitan Regional Council.
Newman, P. and Marks, P., 1979. The removal of heavy metals
by Perth sands. Groundwater Pollution Conference. Perth:
Australian Government Publishing Service, 267–290.
OECD, 1992. Report of the OECD workshop on extrapolation of
laboratory aquatic toxicity data to the real environment.
Environment Monographs No. 59. Paris: Organisation for
Economic Co-operation and Development (OECD).
Pennington, P.L. and Scott, G.I., 2001. Toxicity of atrazine to the
estuarine phytoplankter Pavlava sp. (Prymnesiophyceae):
increased sensitivity after long-term, low-level population
exposure. Environmental Toxicology and Chemistry, 20,
2237–2242.
Peterson, C.G. and Stevenson, R.J., 1990. Post-spate develop-
ment of epilithic algal communities indifferent current
environments. Canadian Journal of Botany, 68, 2092–2102.
Pratt, J.R. and Cairns, J., Jr., 1996. Ecotoxicology and the redun-
dancy problem: understanding effects on community structure
and function. In: M.C. Newman and C.H. Jagoe, eds. Ecotox-
icology: a hierarchical treatment. BocamRaton, FL: Lewis
Publishers, 347–370.
Ratte, H.T., 1999. Bioaccumulation and toxicity of silver com-
pounds: a review. Environmental Toxicology and Chemistry,
18, 89–108.
Rosen, B.H., 1994. Use of periphyton in the development of bio-
cretiria. In: W.S. Davis and T.P. Simon, eds. Biological assess-
ment and criteria: tools for water resource planning. Boca
Raton, FL: Lewis Publishers, 209–215.
Salminen, J., van Gestel, C.A.M., and Oksanen, J., 2001. Pol-
lution-induced community tolerance and functional redun-
dancy in a decomposer food web in metal-stressed soil.
Environmental Toxicology and Chemistry, 20, 2287–2295.
Sample, E.B., et al., 1998. A guide to the ORNL ecotoxicological
screening benchmarks: background, development, and appli-
cation. Oak Ridge, TN: Oak Ridge National Laboratory for
the United States Department of Energy.
Schindler, D.W., 1996. Ecosystems and ecotoxicology: a per-
sonal perspective. In: M.C. Newman and C.H. Jagoe, eds.
Ecotoxicology: a hierarchical treatment. Boca Raton, FL:
Lewis Publishers, 371–398.
Suter II, G.W., et al., 1985. Extrapolating from the laboratory to
the ﬁeld: How uncertain are you? In: R.D. Cardwell, R. Purdy
and R.C. Bahner, eds. Aquatic Toxicology Hazard Assess-
ment: 7th Symposium, ASTM STP 854. American Society
for Testing and Materials, Philadelphia, PA, USA, 400–413.
United States Environmental Protection Agency, 1986. Quality
criteria for water. EPA 440/586–001. Washington, DC:
Unied States Environmental Protection Agency.
United States Environmental Protection Agency, 2000. Stressor
identiﬁcation guidance document. Washington, DC: United
States Environmental Protection Agency.
WAWC, 1994. Harmful phytoplankton surveillance in Western
Australia. Perth: Waterways Commissions.
Weitzel, R.L., 1979. Methods and measurements of periphyton
communittes: a review, ASTM STP 690. Philadephia, PA:
American Society for Testing and Materials.
Yasuno, M. and Whitton, B.A., 1988. Biological monitoring of
environmental pollution. Tokyo: Tokai University Press.

JBRM2012

More Related Content

What's hot (19)

Viewers also liked (19)

Similar to JBRM2012 (20)

JBRM2012