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4. ENVIRONMENTAL SURVEY RESULTS
4.1 SITE DESCRIPTIONS AND HABITAT ASSESSMENT
Table 4-1 Summarised habitat descriptions for each site
Site (representative image) Description
Burlong Pool (control site upstream of WWTP) Burlong Pool is an analogue site located 9.5 km upstream of the WWTP discharge and is of similar size and condition to the potential impact site, Katrine Pool. It is one of the 16 priority rehabilitation pools identified by the DoW (2007). Situated between farmland with highly disturbed riparian vegetation consisting of fragments of native vegetation on right bank and predominantly exotic grasses, and good coverage on the left bank. Local grazing. Highly eroded banks in places as a result of vegetation clearing. Limited shading of the channel. Submerged and emergent macrophytes present (~30% of channel). Small amount of woody debris present. Deep channel in parts and heavily vegetated with macrophytes, although noticeably less after February 2017 floods. A small culvert at the road crossing/bridge allows pool formation as the river dries.
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Site (representative image) Description
Mortlock (Mortlock River upstream WWTP) A control site situated at the edge of Northam township on the Mortlock River, just upstream from the confluence with the Avon River, and assessed for water quality only. The Mortlock River is known to be a significant contributor to nutrient and salt concentrations in the lower Avon River. The site is situated between grazing land and a public recreation/parking area (Enright Park Corroborree Site), with a small cement weir/river walkway presenting a physical barrier to fish passage in some flows. Extremely disturbed riparian area, dominated on the left bank by exotic grasses and providing minimal shading (<5% of channel). Severe bank erosion on the right bank with undercutting. No obvious channel modifications. Small amount of macrophytes present (5% cover). Moderate deposition including vegetated instream bars.
US 1.8km (control site upstream WWTP 1.8km) The control site US 1.8km site represents similar habitat to the potential impact site DS 3.0km and contains inflow from the Mortlock River. This site is representative of conditions immediately prior to the WWTP discharge location.
Site upstream of road crossing/bridge and culvert. Riparian zone cleared in places, dominated by exotic grasses (95%) and Sheoaks (25%), overall of a good width. Some grazing evident. Vegetation overall very highly disturbed including valley clearing. Vegetated mid-channel bars. Low macrophyte presence (5%).
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Site (representative image) Description
WWTP Out Situated on a bend in the Avon River approximately 4.5km downstream of the Northam weir. Water quality only was measured at this site, at a point approximately 150m prior to release into the Avon River.
DS 3.0 (downstream WWTP 3.0km) The potential impact site 3km downstream of the WWTP, and upstream of Katrine Pool, will provide information on impact gradients (if any) and replicate to assess variability within the downstream habitats.
Situated on farmland with local impacts associated with grazing and direct access of stock to water way, and beyond riparian zone, cropping. Overall very high disturbance to vegetation including intrusion of exotic grasses and clearing to valley vegetation. Good riparian coverage of banks but providing limited shading of the channel. Emergent macrophytes present (~10% of reach). Small amount of woody debris present. Exotic grasses present on mid- channel bars.
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Site (representative image) Description
Katrine Pool (downstream WWTP 10km) Katrine Pool, approximately 10 km downstream of the WWTP discharge, has been identified as a high value aquatic habitat, recreational area and local government reserve. This pool is also one of 16 in the Avon Catchment identified by the DoW as a priority for rehabilitation (DoW 2007). The Katrine Pool sampling site has been selected as the key downstream receptor site for this impact assessment.
Local landuse includes recreational picnic areas and cropping. Vegetation highly disturbed with clearing to valley but with native species present and good to excellent coverage in riparian zones. Groundcover predominantly exotic grasses. High percentage cover of macrophytes (40%). Road access point was washed out during the February 2017 floods, though this is unlikely to have impacted greatly on the site.
Table 4-2 Summarised USEPA Habitat characterisation for Low Gradient Streams with total low gradient habitat score (out of a possible 200)
Site bank bank bank bank (N/200)
Epifaunal score Pool variability Pool susbstrate characterization Channel sinuosity channel alteration stability right bank channel flow status sediment deposition Riparian score zone left Bank stability Left bank Bank Riparian score zone right Vegetative protection left substrate/available cover Total low habitat gradient vegetative protection right
Burlong 14 11 6 5 20 15 6 6 3 6 6 7 1 106
Mortlock 6 13 N/A 10 18 15 9 3 1 8 6 3 2 94
US 1.8 6 12 12 5 19 12 7 6 6 6 6 7 8 112
DS 3.0 6 11 11 6 18 13 8 7 7 6 6 7 7 113
Katrine 10 12 14 3 19 13 10 5 7 6 6 5 9 119
4.2 WATER QUALITY
4.2.1 FIELD PARAMETERS The major physico-chemical difference between the TWW discharge water quality and the receiving waters of the Avon River across all sampling seasons is the level of salinity. The Avon River sites ranged from 4.90-18.76 ppt salinity, the Mortlock River was more saline ranging from 18.47-26.75 ppt, whereas the
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WWTP site was fresher than all other sites with salinity values ranging 1.11-1.46 ppt (Figure 4-1). The most saline values at all river sites were recorded in the recessional period, with the least saline values for each site being recorded in the high-flow sampling period. The influence of the WWTP discharge on salinity downstream was negligible with DS 3.0 recording similar salinity levels to US 1.8 across the sampling periods (DS 3.0 = 11.80, 18.52, 14.13 and 13.33; US 1.8 = 11.61, 17.68, 14.14 and 12.07).
While the Avon River is considered mid-high saline (salinity levels recorded for the survey were orders of magnitude above the ANZECC/ARMCANZ (2000) guidance for south-west lowland rivers (120-300 µS/cm)), the water at the WWTP site was marginal-high brackish (DoW 2011a), above fresh water guidance for south-west lowland rivers though within ranges (for example) for moderately salt tolerant irrigation purposes and stock watering (ANZECC/ARMCANZ 2000). The Avon River is well above suitable salinity for irrigation of even very salt tolerant plant species or any stock watering.
The pH of the TWW is typically at the lower end of the range within the Avon River though within the ANZECC/ARMCANZ (2000) guidance for south-west lowland rivers (pH 6.5-8). The lowest pH recorded for the survey was 6.07 at the upstream reference pool (Burlong Pool) during the high-flow period. This site also recorded the lowest pH during the recessional sampling (Figure 4-1). No influence of the TWW discharge on downstream Avon River pH was apparent.
Dissolved oxygen was lower within the TWW discharge than the receiving waters, with elevated DO recorded at several sites during the high-flow and low-flow periods (Figure 4-1). There was no consistent trend with DO downstream of the TWW discharge with DO reducing (upstream to downstream) during the recessional period and increasing (upstream to downstream) during the low-flow period. It was not considered likely that there was a significant impact to DO from the TWW discharge.
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Figure 4-1 Salinity, pH and Dissolved Oxygen by site and season. Note: dissolved oxygen (% saturation) was recorded at Mortlock in the low-flow resample but measured far higher than all other sites at 567.9 so was excluded from the chart for clarity reasons. There was no indication of impacts to water temperature downstream of the TWW discharge (Figure 4-2)
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Figure 4-2 Water temperature by season and site
There is no indication of an impact to dissolved oxygen levels in the Avon River from the Water Corporation long-term monitoring program (2013 to 2017; Figure 4-3). See Figure 3-2 in Section 3.1 for site locations.
Figure 4-3 Dissolved oxygen from Water Corporation long-term monitoring (2013-2017)
4.2.2 LOGGER DATA Two water level, conductivity and water temperature loggers were installed for the study program at the 1.8 km upstream site and Katrine Pool respectively. Unfortunately, the data for these loggers was compromised by the large flood event in February 2017 in which one logger was buried in sediment though still recoverable (upstream 1.8 km) and the other (Katrine Pool) was buried too deeply to be recovered. A data download did occur mid-study in November 2016 which allowed comparison of the upstream and downstream sites for the high-flow to mid-recessional periods. The dataset for the upstream site is considered reasonable until the flood event and therefore provides a record to early February 2017 (Figure 4-4). After this date the water temperature variability drops, which is an indicator that the logger was covered by sediments and therefore the salinity and temperature data is likely to be only partially
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representative of the overlying water. Water pressure however does tend to pass through the sediments to a greater degree such that the water level data beyond this data is likely to be more reliable.
The notable feature of the salinity record is the rapid rise at the later part of the recessional phase, from 22 mS/cm to 57 mS/cm (above seawater levels). An apparent flow event in the third week of December 2016 caused a sharp drop in conductivity back to 32 mS/cm before a steady rise again. This data should be treated with caution as it cannot be verified and the salinity at its peak is significantly higher than generally recorded in the Avon River in this area. Figure 4-5 provides a comparison of the water level data from the logger at the 1.8 km upstream site and the DoW records from the Northam Weir, Toodyay and Mortlock East gauging stations on the Avon River, together with the daily rainfall record at Northam for the summer flood event. It is interesting that the Toodyay and Mortlock sites pick up the general pattern for the flood event seen in the study logger data however, the initial peak is largely absent at Northam. The Mortlock river accounts for the vast majority of the inflow to the Avon River during this initial flood peak, suggesting localised rainfall within that catchment.
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Figure 4-4 Conductivity, water level and temperature logger data from the 1.8 km upstream site FIGURE NOTE: the dark grey line data from September 2016 to November 2016 is what is available for the Katrine Pool site. This logger was lost in the February 2017 floods and with it the November 2016 to April 2017 data.
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Figure 4-5 Comparison of US 1.8 km logger water level with Northam, Toodyay and Mortlock gauging data together with daily rainfall at Northam – January to April 2017
4.2.3 IONIC COMPOSITION The water type at all sites was Sodium-Chloride (Na-Cl) dominated with little variability in ionic composition. The TWW displayed slightly more alkalinity and sulphate content though was still largely on the Na-Cl spectrum (Figure 4-6).
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Major Ions by Season (Piper Diagram)
Legend
80% A BURLONG 80% B Mortlock 60% Ca + Mg D US 1.8 60% + Cl WWTP OUT 4 G 40% SO J DS 3.0 40% I Katrine 20% A Burlong AJD 20% Mortlock GJAIDIB B G B D US 1.8 J DS 3.0 I Katrine A BURLONG B MORTLOCK D US 1.8 20% G WWTP 80% 20% Na + K 80% 3 J DS 3.0 40% KATRINE HCO I SO 60% 40% 60% Mg 4 60%
40% 60% 40% 80% 20% AJDI 80% G 20% GADJIBB G ABIJIBDAJ 80% 60% 40% 20% DA
20% 40% 60% 80% Ca Cl
Figure 4-6 Piper diagram showing major ion distribution by site and season – high-flow (blue), recessional (red) and low-flow (green) The figure above (Figure 4-6) shows the major ion distribution at each site coloured by flow season. Upstream sites are shown with no colour fill while downstream sites have a solid colour fill in the symbol. There was no indication of significant impacts to the major ion composition of the Avon River due to the TWW discharge.
4.2.4 METALS A comprehensive suite of total and dissolved metals and metalloids were measured at all 6 sites. These included aluminium, arsenic, boron, cadmium, chromium, copper, lead, manganese, mercury, nickel, selenium and zinc in both total and dissolved form.
4.2.4.1 DISSOLVED METALS Dissolved aluminium was detected on one occasion during initial sampling at WWTP during the low-flow (April 2017) period at just above the detection limit (0.02 mg/L; LOR 0.01 mg/L) (see Appendix B). This was below the freshwater guideline of 0.055 mg/L in pH>6.5 water (95% EPL; ANZECC/ARMCANZ 2000). During the low-flow resample however, dissolved aluminium was detected at two sites (Katrine Pool and Burlong Pool). Katrine Pool recorded levels of dissolved aluminium at 0.06 mg/L (greater than the ANZECC guideline) with Burlong Pool recording levels of 0.03 mg/L.
Dissolved boron was detected at several sites above the ANZECC guideline (0.37 mg/L; 95% EPL) across all sampling periods. Values ranged from 0.15 mg/L (Burlong Pool, low-flow resample February 2018) to 2.45 mg/L (Mortlock River, low-flow resample February 2018). Dissolved boron was only below the ANZECC guideline on 3 occasions but exceeded the guidelines on 22 occasions across all sampling periods, with values at Mortlock consistently higher than all other sites. The values of US 1.8 and DS 3.0 are similar (the same for 2 sampling seasons with a minor increase at the DS 3.0 site in 2 sampling seasons). This minor increase however is likely due to surrounding land use or other contributors as opposed to the WWTP. The
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value of boron at WWTP out site was consistently lower than the Avon river values immediately above and below the discharge site (as measured at US 1.8 and DS 3.0) (see Figure 4-7).
3 Detection limit = 0.05 mg/L 2.5
2 High-flow 1.5 Recessional Low-flow
Boron Boron (mg/L) 1 Low-flow resample ANZECC 0.5
0 Burlong Mortlock US 1.8 WWTP DS 3.0 Katrine Site
Figure 4-7 Boron measured at sites along the Avon River including ANZECC guideline value and detection limit.
Dissolved copper was only detected above the freshwater guidelines (0.0014 mg/L; 95% EPL) at one site on one occasion (Mortlock River – high-flow; 0.011 mg/L). The dissolved copper value at Mortlock River during the high-flow period remains above the guideline even when modified for hardness (Hardness = 194 mg/L; HMTV4 0.0068 mg/L).
While dissolved manganese and dissolved nickel were detected at various sites across the seasonal sampling, all detections were below guideline levels. Dissolved arsenic, cadmium, chromium, lead, mercury, selenium and zinc were below detection limits in all samples on all dates.
4.2.4.2 TOTAL METALS From the suite of metals and metalloids tested for (see Appendix B), 8 total metals were recorded at values above the limit of reporting throughout the survey (Table 4-3). Of these, aluminium, manganese and boron were detected during every sampling season. Boron exceeded the ANZECC guideline value on all but 3 occasions, with aluminium exceeding the ANZECC guideline value on >50% of sampling occasions. Although manganese was consistently recorded at all sites, the value never breached the ANZECC trigger value. Total copper was detected above the ANZECC guideline values on 4 occasions, twice during the dry sampling in 2017 (US 1.8: 0.002 mg/L, DS 3.0: 0.002 mg/L; ANZECC 95%: 0.0014 mg/L) and twice during the dry re-sample 2018 (Burlong: 0.002 mg/L, Katrine: 0.004 mg/L). When the hardness of each site was taken into consideration however, the values fell below the hardness modified trigger value (HMTV) for 3 of the 4 sites, with Katrine still breaching the HMTV (refer to Table 4-3 for HMTV’s). Chromium was the only other total metal to breach the ANZECC trigger value. This occurred during the dry-resample 2018 (Katrine Pool: 0.0025 mg/L; ANZECC 95% 0.001 mg/L), however, when hardness was taken into consideration the value equalled the hardness modified trigger value (HMTV: 0.0025 mg/L).
Table 4-3 Total metals detected above limits of reporting across the seasons including ANZECC guideline trigger values for reference. Bold values indicate breaches of trigger values.
LOR ANZECC Sampling Variable (mg/L) (95%; mg/L) Season Burlong US 1.8 Mortlock WWTP DS 3.0 Katrine
Aluminium 0.01 0.055 Wet 0.39 0.26 0.23 0.045 0.22 0.24
4 HMTV: Hardness modified trigger value as per Table 3.4.3 and Table 3.4.4 in ANZECC/ARMCANZ (2000)
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LOR ANZECC Sampling Variable (mg/L) (95%; mg/L) Season Burlong US 1.8 Mortlock WWTP DS 3.0 Katrine
Recessional <0.01 <0.01 <0.01 - <0.01 0.1
Dry 2017 0.07 0.05 <0.01 0.01 0.13 0.07
Dry 2018 0.5 <0.01 0.08 - <0.01 1.105
Arsenic 0.001 0.024 Wet <0.001 <0.001 <0.001 <0.001 <0.001 <0.001
Recessional <0.001 <0.001 <0.001 - <0.001 <0.001
Dry 2017 <0.001 <0.001 <0.001 <0.001 <0.001 <0.001
Dry 2018 <0.001 <0.001 0.006 - <0.001 <0.001
Copper 0.001 0.0014 Wet 0.001 0.001 <0.001 <0.001 0.001 0.001
Recessional <0.001 <0.001 <0.001 <0.001 <0.001 <0.001
0.00728* Dry 2017 <0.001 0.002 <0.001 <0.001 0.002 <0.001
0.0035a Dry 2018 0.002 <0.001 <0.001 - <0.001 0.004
Lead 0.001 0.0034 Wet <0.001 <0.001 <0.001 <0.001 <0.001 <0.001
Recessional <0.001 <0.001 <0.001 - <0.001 <0.001
Dry 2017 <0.001 <0.001 <0.001 <0.001 <0.001 <0.001
Dry 2018 <0.001 <0.001 <0.001 - <0.001 0.001
Manganese 0.001 1.9 Wet 0.026 0.031 0.035 0.039 0.029 0.031
Recessional 0.017 0.292 0.07 - 0.048 0.214
Dry 2017 0.133 0.178 0.132 0.138 0.207 0.231
Dry 2018 1.18 0.242 0.403 - 0.257 0.595
Chromium 0.001 0.001 Wet <0.001 <0.001 <0.001 <0.001 <0.001 <0.001
Recessional <0.001 <0.001 <0.001 - <0.001 <0.001
Dry 2017 <0.001 <0.001 <0.001 <0.001 <0.001 <0.001
0.0025^ Dry 2018 <0.001 <0.001 <0.001 - <0.001 0.0025
Nickel 0.001 0.011 Wet 0.001 0.002 <0.001 <0.001 0.002 0.002
Recessional <0.001 <0.001 <0.001 <0.001 <0.001 <0.001
Dry 2017 0.002 0.002 <0.001 <0.001 0.003 0.003
Dry 2018 0.002 <0.001 <0.001 - <0.001 0.004
Boron 0.05 0.37 Wet 0.39 0.56 1.38 0.47 0.56 0.59
Recessional 0.59 0.77 1.98 - 0.84 0.79
Dry 2017 0.85 0.97 2.00 0.63 0.99 1.045
Dry 2018 0.14 0.67 2.46 - 0.83 0.245
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*HMTV for Cu based on Hardness category of “Very Hard (180-240)” as measured at US 1.8 and DS 3.0 during Dry 2017 aHMTV for Cu based on Hardness category of “Hard (60-119)” as measured at Burlong and Katrine during Dry 2018 ^HMTV for Cr(III) based on Hardness category of “Hard (60-119)” as measured at Katrine during Dry 2018
There was no indication of any impact to dissolved or total metal levels downstream of the TWW discharge point in the Avon River.
4.2.5 HYDROCARBONS Total petroleum hydrocarbons (TPH), BTEX (benzene, toluene, ethyl-benzene and xylene) and naphthalene were sampled at one upstream (US 1.8) and one downstream site (DS 3.0) during the initial high-flow and recessional sampling, and during the low-flow resample. No detections were made at any site during initial sampling, however, during the low-flow resample in February 2018 several detections of total petroleum hydrocarbons were made (see Appendix B)
The C15 – C28 Fraction was recorded above the limit of reporting (<100 µg/L) at DS 3.0 (120 µg/L) and US 1.8 (330 µg/L). The C29 – C36 Fraction was also recorded above the limit of reporting (<50 µg/L) at US 1.8 (50 µg/L). Total recoverable hydrocarbons were also recorded at both sites during the low-flow resample. The >C16 – C34 fraction was detected above the limit of reporting (<100 µg/L) at both the DS 3.0 site (140 µg/L) and US 1.8 site (330 µg/L).
4.2.6 SURFACTANTS Anionic surfactants as Methylene blue active substances (MBAS) were sampled during the high-flow period at sites upstream (US 1.8) and downstream (DS 3.0 and Katrine Pool) of the WWTP discharge as well as the WWTP itself. All samples during the initial survey were below detection limits at <0.1 mg/L MBAS (see Appendix B).
During both low flow periods, April 2017 and the low-flow resample in February 2018, Anionic surfactants as MBAS were only sampled at one upstream site (US 1.8) and one downstream site (DS 3.0). All samples during the initial low-flow survey were below detection limits (<0.1 mg/L). However, during the low-flow resample detections of anionic surfactants as MBAS were made at both DS 3.0 (0.3 mg/L) and US 1.8 (0.2 mg/L). Both of which were either higher than or equal to the ANZECC guideline for freshwater for recreational purposes (0.2 mg/L; ANZECC/ARMCANZ 2000).
4.2.7 HERBICIDES AND PESTICIDES A total of 41 Orthophosphate (OP) Pesticides, 12 Thiocarbamates and Carbamates, 2 Dinitroanilines, 2 Triazinone Herbicides, 10 Conazole and Aminopyrimidine Fungicides, 5 Phenylurea, Thizdiazolurea, Uracil and Sulfonylurea Herbicides, 1 Metolachlor, 9 Triazine Herbicides and 5 miscellaneous pesticides were tested for in samples taken during the high-flow period at one upstream site (US 1.8) and one downstream site (DS 3.0). No OP pesticides or fungicides were detected (see Appendix B).
Three herbicides were detected at just above the detection limit at both the upstream and downstream sites (Table 4-4). These were Diuron (US 1.8: 0.05 µg/L, DS 3.0: 0.03 µg/L; LOR <0.02 µg/L), Fluometuron (US 1.8: 0.01 µg/L, DS 3.0: 0.02 µg/L; LOR <0.01 µg/L) and Metolochlor (US 1.8: 0.02 µg/L, DS 3.0: 0.02 µg/L; LOR <0.01 µg/L).The herbicide Simazine was only detected at the upstream site (0.06 µg/L; LOR <0.02 µg/L) at well below the ANZECC guideline (3.2 µg/L; ANZECC/ARMCANZ 2000).
Table 4-4 Herbicides detected above limits of reporting, comparing US 1.8 to DS 3.0 values and ANZECC guidelines where available
Sampling Variable LOR (µg/L) ANZECC (95%; µg/L) Season US 01.8 DS 3.0
Diuron 0.02 ID* Wet 0.05 0.03
Fluometuron 0.01 - Wet 0.01 0.02
Metolochlor 0.01 100a Wet 0.02 0.02
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Sampling Variable LOR (µg/L) ANZECC (95%; µg/L) Season US 01.8 DS 3.0
Atrazine 0.01/0.5^ 13 Wet 0.08 0.52
Dry 2018 <0.5 <0.5
Simazine 0.02 3.2 Wet 0.06 <0.02 *ID – Insufficient data a maximum concentration based on recreational purposes ^ Note: LOR for the initial high-flow sampling of Triazine Herbicides was 0.01 µg/L for Atrazine. The individual test for Atrazine in the dry re-sample has a LOR of 0.5 µg/L.
Atrazine (herbicide) was detected at both sites, with a marked increase from upstream to downstream, from 0.08 µg/L at US 1.8 to 0.52 µg/L at DS 3.0 (despite this increase both records fell below the guideline value of 13 µg/L; ANZECC/ARMCANZ 2000). Atrazine was re-tested for during the low-flow resample (February 2018) at two sites, US 1.8 and DS 3.0. Atrazine was below detectable limits (LOR 0.5 µg/L) at both sites during this sampling occasion.
4.2.8 NUTRIENTS The nutrient levels within the Avon River and the Mortlock River in particular are known to be generally elevated due to catchment land use practices including agricultural run-off and WWTP discharges (Hennig and Kelsey 2015). The water sampling undertaken for the current study looked at a range of nutrient parameters including ammonia, nitrite, nitrate, total kjeldahl nitrogen, total nitrogen, total phosphorus and reactive phosphorus (a proxy for phosphate). While the TWW was elevated in most of the nutrient parameters, there was little indication of an impact on the receiving waters of the Avon River (Figure 4-8 and Figure 4-9).
TN was generally slightly above the ANZECC/ARMCANZ (2000) guidance for south-west lowland rivers (1.2 mg/L) at both upstream and downstream sites, with the highest recording (other than the WWTP out; 28 mg/L) noted from the Mortlock River in the low-flow resample at 17.2 mg/L. Ammonia was equal to or above the guidelines (0.08 mg/L) during the recessional period at four sites. During the high-flow and initial low-flow sampling however, ammonia was below these levels at all sites (excluding WWTP out). During the low-flow resample ammonia was recorded above the guideline levels at both Burlong Pool (0.47 mg/L) and Katrine Pool (0.13 mg/L) but not at the remaining sites. Nitrate+nitrite (NOx) was below ANZECC guidelines (0.15 mg/L) at all sites other than the WWTP out during the initial sampling. However, during the low-flow resample, Nitrate+nitrite (NOx) was above the guideline level at both Burlong Pool (2.51 mg/L) and Katrine Pool (2.51 mg/L).
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100 1
Detection limit Detection limit 10 = 0.01 mg/L = 0.01 mg/L
1 0.1
0.1
0.01 0.01 Nitrite Nitrite as N (mg/L) Ammoniaas N (mg/L)
0.001 0.001
High flow Recessional Low-flow Low-flow resample ANZECC High flow Recessional Low-flow Low-flow resample
10 10 Detection limit Detection limit
= 0.01 mg/L = 0.01 mg/L
1 1
(mg/L) 0.1
0.1 x
NO 0.01 0.01 Nitrate Nitrate as N (mg/L) 0.001 0.001
High flow Recessional Low-flow Low-flow resample High flow Recessional Low-flow Low-flow resample ANZECC
100 100 Detection limit Detection limit = 0.1 mg/L = 0.1 mg/L 10 10
1 1 0.1 0.1
Total N (mg/L) 0.01 0.01 0.001 0.001 Total Kjeldahl Nitrogen as N (mg/L) High flow Recessional Low-flow Low-flow resample High flow Recessional Low-flow Low-flow resample ANZECC
Figure 4-8 Nitrogen species concentration from upstream to downstream by site and flow season including detection limits and ANZECC guidelines where applicable.
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TP ranged from not detected (<0.01 mg/L) to 0.16 mg/L during the initial sampling periods (see Appendix B). The values were generally either just below or slightly above the ANZECC/ARMCANZ (2000) south-west lowland river guidelines (0.065 mg/L) in upstream and downstream sites for all initial sampling periods. During the low-flow resample, the TP values were notably higher with all sites exceeding the guideline value, including Mortlock River which recorded the highest value of all sites (1.94 mg/L). There did not appear to be any impact from the WWTP for total P, with US 1.8 and DS 3.0 recording similar values during the low-flow resample (0.16 mg/L and 0.20 mg/L, respectively). Filterable reactive phosphorus was below the ANZECC guidelines (0.04 mg/L) at all sites except the Mortlock River and WWTP out.
10 1 Detection limit Detection limit = 0.01 mg/L = 0.01 mg/L 1 0.1
0.1
0.01
Total P (mg/L) 0.01
0.001 0.001 Reactive Reactive Phosphorus as P (mg/L)
High flow Recessional High flow Recessional Low-flow Low-flow resample Low-flow Low-flow resample ANZECC ANZECC
Figure 4-9 Phosphorus species concentration from upstream to downstream by site and flow season, including detection limits and ANZECC guideline values. The pattern of nutrient distribution upstream and downstream of the TWW discharge point was largely mirrored in the long-term water quality monitoring dataset provided by the Water Corporation as part of their compliance monitoring program. The closest site to the TWW discharge point in the Water Corporation monitoring program is only 700 m downstream and on the same bank as the discharge point (see Figure 3-2 in Section 3.1 for site locations). This is likely to lead to sampling of poorly mixed flows containing elevated levels of TWW, particularly during lower river flow periods. There were slightly elevated nutrients in all parameters measured from 1.0 km upstream to 0.7 km downstream of the TWW discharge point, however, this had mixed to background concentrations by 1.5 km downstream (Figure 4-10 – Ammonia; Figure 4-11 – Nitrate; Figure 4-12 – Total Nitrogen and Figure 4-13 – Total Phosphorus). Figure 4-12 showing Total Nitrogen ranges has been overlain with the ANZECC and FARWH guidance values for reference. It can be seen that by 1.5 km downstream the median TN concentration was below the 80th percentile of the upstream site, indicating compliance with ANZECC site specific guidance at this location.
Figure 4-14 shows the Total Phosphorus statistical ranges by season from the Water Corporation long term monitoring program. It can be seen from this figure that the pattern of upstream and downstream distributions was similar across the seasons though with slightly higher values during the summer low flow period.
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Figure 4-10 Statistical ranges of ammonia in Water Corporation long-term monitoring data (2013-2017)
Figure 4-11 Statistical ranges of Nitrate in Water Corporation long-term monitoring data (1996-2017)
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Figure 4-12 Statistical ranges of Total Nitrogen in Water Corporation long-term monitoring data (1996- 2017)
Figure 4-13 Statistical ranges of Total Phosphorus in Water Corporation long-term monitoring data (1996- 2017)
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Figure 4-14 Statistical ranges by season of Total Phosphorus in Water Corporation long-term monitoring data (1996-2017)
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4.2.9 MICROBIOLOGY The microbiology sampling returned mixed results with respect to any potential impacts to the Avon River. While levels of faecal coliforms and enterococci were higher on occasion downstream of the TWW discharge, this was not consistent or linear with distance downstream. During the high-flow sampling both microbiology parameters were similar or lower at 3.0 km downstream when compared to the 1.8 km upstream site. However, levels increased again at Katrine Pool, possibly due to the larger size of the pool, lower flow velocities, sediment contribution and potential run-off from the surrounding catchment (DoW, 2008). For the recessional phase the faecal coliforms at 3.0 km downstream were elevated (110 cfu/100ml), but then reduced to below US 1.8 levels (40 cfu/100ml) by Katrine Pool (24 cfu/100ml) (Figure 4-15).No pattern was observed for enterococci during the recessional period with US 1.8 recording the highest value (17 cfu/100ml) and Burlong Pool, DS 3.0 and Katrine Pool all recording similar values (4 cfu/100ml, 5 cfu/100ml and 5 cfu/100ml, respectively). The Katrine Pool sample for the low-flow (refugia) sampling phase did not pass quality control but data from the 1.8 km upstream to 3.0 km downstream site showed similar levels of bacteria upstream and downstream of the TWW discharge (Figure 4-15).
During the low-flow resample, both Burlong Pool (2300 cfu/100ml) and Katrine Pool (38,000 cfu/100ml) recorded values well above the ANZECC guidelines for faecal coliforms for primary contact (150 cfu/100ml) as well as for secondary contact (1000 cfu/100ml). Records for enterococci also exceeded the primary and secondary contact guidelines (35 cfu/100ml and 230 cfu/100ml, respectively) at three sites; US 1.8 (330 cfu/100ml), DS 3.0 (360 cfu/100ml), and Katrine Pool (2300 cfu/100ml). Again, there does not appear to be any immediate impact from the WWTP as both the “reference” US 1.8 site and “impact” DS 3.0 site recorded similar values.
100000 10000
10000 1000 1000 100 100
10 10 1 No No data 1 No data 0.1 0.1 Burlong US 1.8 DS 3.0 Katrine Burlong US 1.8 DS 3.0 Katrine Enterococci (CFU/100mL) Site Site Faecal Faecal Coliforms(CFU/100mL) High-flow Recessional High-flow Recessional Low-flow Low-flow resample Low-flow Low-flow resample
Figure 4-15 Faecal coliforms (left) and Enterococci (right) upstream and downstream of the TWW discharge. Note the log scale due to high values recorded at Katrine Pool in the low-flow resample. While the discharge from the WWTP is likely to be adding a bacteriological loading to the Avon River, the presence of faecal coliforms and enterococci at similar orders of magnitude both above and below the discharge point indicates that other sources are dominant in this section of the river. The change from fresh to saline conditions as the TWW enters the river is likely to cause some die-off of the bacteria in the TWW (e.g. Oren and Vlodavsky 1985). Other sources of bacteriological pathogens in the Avon River include run-off from farming areas, stockyards, feedlots, piggeries and to a lesser degree town septic tanks (e.g. Hennig and Kelsey 2015). In addition to the pathogens introduced to the Avon River through wastewater discharges and run-off, there will be a natural in-situ population of bacteria which will vary in abundance based on available resources and habitats.
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4.3 SEDIMENT QUALITY Samples were taken for sediment quality at 5 sites in total; 3 upstream (Burlong, Mortlock and US 1.8) and 2 downstream (DS 3.0 and Katrine). Total N ranged from 70 mg/kg (Mortlock; initial dry sampling) to approximately 1536.67 mg/kg (Katrine; dry re-sample). Total P ranged from 13 mg/kg (Burlong; wet) to approximately 215 mg/kg (Katrine; dry re-sample). Total Organic Carbon (TOC) ranged from 0.04% (Mortlock; recessional & dry) to 3.09% (DS 3.0; wet).
The dry resample recorded higher values at all sites for TN and TP (apart from US 1.8 which recorded its highest Total N value in the initial dry sampling). During the dry-resample it also appears that TN and TP increase downstream of the WWTP. Distinguishing the cause of this however will be difficult, due to several tributaries feeding into the Avon River between the WWTP and the downstream sites. For example, Mistake Creek feeds into the Avon River less than 1 km upstream from Katrine Pool and has been previously identified as a significant contributor of sediment to Katrine Pool (DoW, 2008). Hampton Brook feeds into the Avon River less than 1 km upstream from DS 3.0 and has also been identified as a source of sediment into the Avon River (DoW, 2008). The pattern was also not consistent, with US 1.8 recording higher values than those sites downstream (DS 3.0 and Katrine) during the initial dry season. There were no consistent indications that sediment TN, TP or organic carbon was higher in sediments below the WWTP in comparison to upstream sediments. In two of the four sampling rounds the highest values were observed in upstream sites for all three parameters.
No pesticides, fungicides, and herbicides that were tested for in samples taken during the recessional period were detected above their respective limits of reporting.
1800 1600
1400 1200 1000 Recessional 800 Dry 600 Wet Total N (mg/kg) 400 Dry resample 200 0 Burlong Mortlock US 1.8 DS 3.0 Katrine Site
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250
200
150 Recessional Dry 100 Wet
Total P (mg/kg) Dry resample 50
0 Burlong Mortlock US 1.8 DS 3.0 Katrine Site
3.5
3
2.5
2 Recessional Dry 1.5 Wet 1 Dry resample 0.5 Total Organic Carbon (%)
0 Burlong Mortlock US 1.8 DS 3.0 Katrine Site
Figure 4-16 Total N, Total P and Total Organic Carbon in sediment
4.4 MODELLING WATER BALANCE
4.4.1 TWW DISCHARGE RATES The Northam WWTP typically discharges to the Avon River during the winter months while there is flow in the river. The flow in the Avon River is highly seasonable and variable between years (Figure 4-17) such that water balance results should be interpreted with the und erstanding that conditions for any given month will also be variable from year to year and month to month. Figure 4-18 provides a timeline of the discharge volumes of TWW to the Avon River from January 2016 to April 2017 with Figure 4-19 displaying the discharge volumes of TWW from October 2017 to April 2018 (during the dry re-sample period). TWW Discharge was seasonal with summers typical having the lowest median daily discharge; worth noting however was the summer of 2016/2017 which despite having a low median had the largest maximum daily discharge during the study period (Table 4-5). The TWW discharge in February to April 2017 was considered unusual and was due to the high rainfall and flood event during February resulting in flood waters entering the sewer system.
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Figure 4-17 Total monthly flow ranges at the Northam Weir – 1970 to 2017
Figure 4-18 Discharge from the Northam WWTP to the Avon River – January 2016 to April 2017 showing sampling dates (red symbols)
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Figure 4-19 Discharge from the Northam WWTP to the Avon River – October 2017 to April 2018 showing dry re-sample period in red
Table 4-5 TWW daily discharge (kL/d) statistics
Summer Spring (Jan/Feb) Autumn Winter Spring Summer (Oct/Nov) Summer Statistic 2016 2016 2016 2016 2016/2017 2017 2017/2018 median 6.448 609.05 525.08 666.56 31.95 398.85 0.06
80th 260.92 1032.13 1080.59 1030.69 921.51 937.22 48.67
20th 0.00 205.49 36.88 30.53 0.00 34.91 0.0 min 0.00 0.00 21.98 0.40 0.00 0.00 0.0 max 1412.05 2254.23 1997.06 1312.51 6997.90 1048.37 1114.9
4.4.2 HYDROLOGICAL LOADING TO THE AVON RIVER Calculations of loading to the Avon River were undertaken using the combined daily flows recorded at the Northam Weir, Mortlock River – Odriscolls Farm, and Mortlock River North gauging stations (representing total flow prior to the WWTP), in comparison to the daily flow on the same date from the WWTP to the Avon River. For days where no flow was recorded at any of the three gauging stations, a flow of 0.45 ML/d was used to take into account the existing water volume in the river between the discharge point and Katrine Pool. This value was derived on a conservative basis using the maximum amount of recorded days without flow in the Avon River (Jan 2016 to April 2017) while the TWW was being discharged (32 days). A conservative estimated volume of river over the same distance (5 m width by 0.3 m depth by 10 km = 15,000 m3 = 15 ML) was divided by the 32 days of discharge to assume that the river had a daily flow/dilution capacity equal to 0.47 ML/d, which was rounded to 0.45 ML/d to be conservative. Under most flow conditions in the Avon River the discharge of TWW comprises a small percentage (median of 0.22%; Figure 4-20). The peak value was 64%, which occurred in March 2016, with similar values reached during the following two dry seasons.
It is important to note that while water balance calculations used WWTP discharge volumes as measured at the final retention pond release point, during the dry season the traverse of the flow through a drain and small wetland (formed by the discharge) mitigates the discharge reaching the Avon River through infiltration and evaporation. All estimates of discharge to the river are likely to be conservative, particularly during the summer months.
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Figure 4-20 Percentage of TWW in the Avon River flow – April 2016 to March 2018 (showing sampling dates – red symbols) There is a distinct lowering of the percentage of TWW in the Avon River during the high-flow season, as would be expected, although this is also the peak TWW discharge period. This pattern is seen more clearly in the 30 day running average of TWW percentage in the Avon River (Figure 4-21). On a 30 day average, which is more likely to represent true river conditions including dilution from existing river volume, the peaks in the early and late part of the season are apparent. The peak value falls to approximately 25% on the 30 day average basis.
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Figure 4-21 Percentage of TWW in the Avon River flow – 30 day running average – April 2016 to March 2018 (showing sampling events – red symbols) It should be noted that the periods where the highest TWW percentage occurs also tend to be periods of elevated salinity (early first flush and end of the recessional period). TWW may provide some relief from these conditions over these periods however, due to the small percentage of TWW in comparison to the receiving Avon River volume this effect would be minimal. As noted above, the secondary peak in February to March 2017 was unusual due to the high rainfall and flooding over this period. The TWW would not normally be discharged over this period due to demand from the council for re-use.
4.5 AQUATIC ECOLOGY
4.5.1 MACROINVERTEBRATES Table 4-6 presents the macroinvertebrate families recorded in each survey period. A total of 20, 23 and 21 families were collected during the peak, recessional, and initial low flow surveys respectively, with 34 unique families present across all initial surveys combined. The low flow re-sample in February 2018 identified 20 families, 5 of which were unique from the earlier surveys, bringing the total unique families/sub-families across the entire survey period to 39.
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Table 4-6 Macroinvertebrate taxa richness for each survey
February Class/Order Family/Sub-family August 2016 November 2016 April 2017 2018
Acarina Acarina x x x x
Odonata Aeshnidae x
Amphipoda Amphipoda x x
Diptera Ceratopogonidae x x x x
Amphipoda Chiltoniidae x x x x
Diptera Chironomidae x
Diptera Chironominae x x x x
Odonata Coenagrionidae x x x
Coleoptera Coleoptera x
Crustacae Copepoda sp. x
Odonata Corduliidae x x x
Hemiptera Corixidae x x x
Lepidoptera Crambidae x
Decapoda Decapoda x
Diptera Diptera x
Coleoptera Dytiscidae x x x x
Coleoptera Elmidae x
Diptera Ephydridae x x
Gastropoda Gastropoda x x
Coleoptera Hydraenidae x x x
Gastropoda Hydrobiidae x x
Coleoptera Hydrochidae x
Coleoptera Hydrophilidae x x x x
Bivalvia Hyriidae x x x
Trichoptera Leptoceridae x x
Odonata Lestidae x x
Odonata Libellulidae x
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February Class/Order Family/Sub-family August 2016 November 2016 April 2017 2018
Gastropoda Lymnaeidae x
Hemiptera Micronectidae x x (Corixidae)
Diptera Muscidae x
Hemiptera Notonectidae x x
Oligochaeta Oligochaeta x x x
Diptera Orthocladiinae x
Crustacae Ostracoda sp. x
Decapoda Palaemonidae x x x
Diptera Stratiomyidae x x
Diptera Tabanidae x
Diptera Tanypodinae x x x x
Gastropoda Tateidae x
Odonata Zygoptera x
Total 20 23 21 20
Macroinvertebrate indices, based on presence absence data, were calculated from sweep samples to provide an indication of site and river condition. Taxa richness, PET taxa richness, % tolerant taxa and % sensitive taxa are calculated and presented in Table 4-7. SIGNAL 2 scores for sweep samples at sites across each season are presented in Table 4-8. NB. The two 'channel' sites (US 1.8 and DS 3.0) in Aug, and DS 3.0 in April, were not suitable for channel sampling due to strong flow during these surveys; both Burlong and Katrine pools had backwaters that enabled channel sampling for these seasons.
Table 4-7 Macroinvertebrate indices
Burlong Pool US 1.8 DS 3.0 Katrine Pool
Aug Nov Apr Feb Aug Nov Apr Feb Aug Nov Apr Feb Aug Nov Apr Feb ‘18 * ‘18 * * ‘18 ‘18 Index
Taxa 11 15 13 12 7 20 17 9 12 12 11 15 11 12 18 12 richness
PET 0 0 1 1 0 0 1 0 0 0 1 1 0 0 1 0 richness
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Index Burlong Pool US 1.8 DS 3.0 Katrine Pool
Tolerant 8 10 9 5 7 14 8 4 9 8 6 8 7 9 10 7 taxa
% tolerant 73 67 69 42 100 70 47 44 75 67 55 53 64 75 56 58 taxa
% sensitive 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 taxa
*Edge only
Table 4-8 SIGNAL 2 scores for edge and channel sampling by site and survey period
Burlong Pool Upstream 1.8 Katrine Pool DS 3.0
Edge Chann Edge Chann Edge Chann Edge Chann el el el el Survey Index
Aug Signal score 2.69 2.79 2.29 NA 2.89 2.83 2.71 NA
Taxa richness 11 6 7 NA 9 8 12 NA
Nov Signal score 2.80 3.19 2.97 2.97 2.81 2.93 3 2.84
Richness 11 12 17 15 12 7 12 8
Apr Signal score 3.24 3.29 3.40 3.34 3.24 3.46 3.13 NA
Richness 11 11 14 12 16 10 11 NA
Feb Signal score 3.45 3.55 3.38 4 3.36 2.8 3.45 3.7 2018
Richness 12 10 8 3 11 5 11 10
All sites had a high percentage (42-100%) of tolerant taxa (SIGNAL grades 1, 2 and 3) across the seasons, with no single sensitive species (SIGNAL grades 8, 9 and 10) recorded. Only one PET order (family Leptoceridae from the order Trichoptera) was collected during the four surveys, which was present at all sites in April 2017 and two sites during the low-flow resample in February 2018. All sites scored SIGNAL 2 Scores of <4 which is indicative of disturbed conditions both upstream and downstream of the WWTP discharge. Figure 4-22 and Figure 4-23 present taxa richness by site and season for the four surveys.
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25
20
15
10
Taxanomicrichness 5
0 Burlong US 1.8 DS 3.0 Katrine Site
High Recessional Low Low-flow resample
Figure 4-22 Macroinvertebrate taxa richness (at family level) by site from upstream to downstream of the WWTP
25
20
15
10
5 TaxanomicRichness
0 High flow Recessional Low flow Low-flow resample Sampling period
Burlong US 1.8 DS 3.0 Katrine
Figure 4-23 Macroinvertebrate taxa richness (at family level) by flow period Edge sampling yielded greater taxanomic richness across each sampling period, apart from Burlong Pool during recessional sampling (see Figure 4-24 for February 2018 taxa richness highlighting this trend).
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14
12
10
8
6 Channel Edge 4 TaxanomicRichness 2
0 Burlong US 1.8 DS 3.0 Katrine Site
Figure 4-24 Macroinvertebrate taxa richness (at family level) by site and sweep location during February 2018 (low-flow resample) Macroinvertebrate communities for all sites and surveys were dominated by salt tolerant species, with no sensitive species recorded, and were low in PET taxa. These findings are indicative of disturbed conditions both upstream and downstream of the WWTP discharge. Macroinvertebrate fauna of south-west Australian streams is known to be depauperate compared to eastern Australia which includes representatives of the PET groups (Halse et al. 2007). Pinder et al. (2004) calculated an average of 40 macroinvertebrate species for Wheatbelt wetlands, so the taxa count of 39 families from the Avon River, (which is likely to represent a greater number of species), suggests a relatively good richness. Mild disturbance has been shown to occasionally increase the number of families at a site (Townsend and Scarsbrook 1997; Halse 2007). The WWTP discharge shows no adverse effects on the salt-tolerant macroinvertebrate communities present.
4.5.2 DIATOMS Twenty-four diatom samples were initially submitted for analysis, consisting of two samples from each site (two sites above and two sites below the WWTP), across three sampling rounds (Aug-Sept 2016; Nov 2016; and April 2017). An additional eight diatom samples were submitted for analysis to capture the low- flow resample in February 2018. Two slides from each of the samples were examined under low and high- power microscopy. The diatom communities above and below the discharge site were analysed in terms of the known ecological preferences and tolerance of diatom species. As discussed in Sections 2.1.3, 2.2.3 and 4.2, the Avon River in Northam has become saline and eutrophic as a result of secondary salinisation and eutrophication due to agriculture and clearing of native vegetation. This has had an impact on the diatom assemblages in the system.
The following details findings for each sampling season. Table 4-9 presents the most common diatom species identified, their relative abundance, and their ecological and biological characteristics.
High-flow (winter) 2016
During this period treated wastewater from the Northam WWTP was discharging in varying volumes into the stream. Altogether a total of 30 diatom species were selected for analysis according to their numerical relative abundance. Every diatom community analysed was dominated only by one or more, often up to five species. During the high-flow period, there was little difference between upstream sites and downstream sites, being dominated by the same species, Tabularia tabulata (see Figure 4-25), a well- known saline species very common in estuaries and saline lakes of Western Australia.
Recessional (early summer) 2016
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During the recessional period in 2016, Burlong Pool, the farthest upstream reference site developed its own diatom community, different from the sites immediately above and below the discharge sites. Similarly, the farthest downstream site, Katrine, developed another unique diatom community. Both these two communities were dominated by diatoms known to be indicators of high salinity (Euhalobous) common in many salt lakes of South Western Australia (Amphora coffeaeformis and Mastogloia species in Burlong Pool and Achnathes brevipes and Mastogloia species in Katrine site). During this time the diatom communities immediately above and below the discharge sites were dominated by Cocconeis placentula, a diatom well known to be common in fresh water tolerant to salinity fluctuations. Thus, there was no impact caused by the discharge that can be detected by diatom communities. The diatom communities in downstream Katrine site seem to free from any impact from discharge having developed its own community of salt tolerant species.
Low-flow (early autumn) 2017
During the autumn of 2017, the dynamics of the diatom communities change further resulting in both Burlong Pool and Katrine site developing their own communities dominated by different species, but all reflecting high salinity. Burlong pool was dominated by Amphora coffeaeformis and Rhopalodia musculus and the downstream site Katrine by Rhopalodia species and Amphora species respectively. Both A. coffeaeformis and R. musculus are known to be highly tolerant of saline environments (Blinn et al., 2004). For the first time we see a noticeable difference in community structure between the immediate upstream and downstream sites (see Figure 4-26). The upstream site was dominated by Amphora coffeaeformis, Rhopalodia and Mastogloia species. The downstream site was dominated by Tabularia tabulata- the species that dominated the system during the winter in 2016. These two communities are showing some differences but both represent eutrophic saline communities.
Low-flow resample (February) 2018
During the summer (low flow) re-sample, a total of 72 diatom species were recorded, however, the diatom communities at all sites were dominated by a single species Amphora veneta (also known as Halamphora veneta). This species alone accounted for >30% of the diatom abundance at all sites. Also dominant in the system during this time were Cyclostephanos sp., Cyclotella meneghiniana (see Figure 4-27), Cyclotella pseudostelligera and other Cyclotella sp. There was little difference in community composition between US 1.8 (25 species) and DS 3.0 (26 species), sharing 12 common species and having very similar percentage of abundances for dominant species (Figure 4-28). Katrine Pool recorded 26 species, 10 of which were also recorded in US 1.8 and DS 3.0, and 9 of which were recorded at Burlong Pool. Burlong Pool appears the most different compared to the remaining sites during this sampling period, including holding the highest species richness (34 species); despite still being dominated by Amphora veneta and Cyclotella sp. Sensitivity values were scored during this round and are presented in Figure 4-29. Sensitivity scores range from 1-100 with 100 indicating highly sensitive species that would likely be found in pristine, non-polluted environments, and 1 indicating highly tolerant species that would survive a range of external stressors (Chessman et al., 2007). All sites scored in the mid-range indicating communities composed of relatively tolerant species.
Table 4-9 List of the most common diatom species identified during this study, their relative abundance, and their ecological and biological characteristics
Ecological and biological Species characteristics Abundance
Achnanthes brevipes Common in estuaries; indicator of Dominant in Katrine in initial summer salinity sampling, but only in small numbers in summer re-sample
Amphora Salinity indicator; salinity low to high Dominant upstream in summer and at coffeaeformis Katrine in autumn
Amphora holsatica High salinity indicator Subdominant in winter in upstream sites
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Ecological and biological Species characteristics Abundance
Amphora veneta Lives in eutrophicated waters Dominant in summer re-sample (also Halamphora veneta)
Bacillaria paxillifer Indicator of high nutrients Common in winter upstream
Cocconeis placentula Freshwater species tolerant to some Most dominant in initial summer salinity sampling in the up/down- stream sites. Present in all sites (minus Katrine Pool) in low numbers during summer re-sample.
Conticribra Indicator of low to high salinity Subdominant in winter weissflogii
Cyclostephanos sp. Found in high abundance in both Found in large numbers across all sites eutrophic lakes and brackish systems during the summer re-sample
Cyclotella Found in high-electrolyte water bodies Common throughout meneghiniana
Cyclotella Common in nutrient rich systems Common across all sites (other than pseudostelligera Burlong Pool) in the summer re-sample
Diploneis Common in marine waters Common in most sites chersonensis
Diploneis ovalis Indicator of low to high salinity Common in most sites
Discostella stelligera Both in lakes and rivers, responds to Found across all sites during summer re- increased nutrients sample
Gyrosigma sp Benthic species indicator of shallow Common in winter water bodies
Mastogloia halophila ‘Salt loving’ species Common in autumn
Mastogloia pumila Salt water species Common in autumn
Mastogloia reimerii Salt water species Common in autumn
Melosira Estuarine species Common in winter nummuloides
Melosira octogona Salt tolerant species Common in winter
Navicula elegans Common in hypersaline waters Common in up/down- stream sites
Navicula salinarum One of the best indicators of salinity Common in most sites and eutrophication
Navicula tripunctata Estuarine and marine Common in winter
Nitzschia hummii Indicator of high salinity and high Subdominant in winter in upstream and nutrients down stream
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Ecological and biological Species characteristics Abundance
Nitzschia linearis Common in both freshwater and Found across all sites in the summer re- estuarine waters sample
Planothidium Tolerant to high nutrients Common in all sites delicatulum
Planothidium Estuarine species Common in upstream sites hauckianum
Planothidium Tolerant to high electrolyte water Common in most sites lanceolatum bodies
Pleurosigma Salt water species Observed at all sites apart from Burlong elongatum Pool during the summer re-sample
Rhopalodia gibberula Well-known as a salt-tolerant species Subdominant in autumn in downstream sites
Rhopalodia musculus Well-known as a salt-tolerant species Dominant in Burlong in autumn
Tabularia tabulata Well known for its tolerance to wide Most dominant species in the system range of salinity and abundant in during initial sampling, however, in lesser shallow streams with high nutrients numbers during summer re-sample and salinity
Tryblionella Common in shallow estuarine systems Common in all sites hungarica
Moreover, the salt and nutrient-tolerant species Tabularia tabulata is much larger in size (biomass) than all other diatoms observed. Therefore, their numerical dominance gives us the most conservative representation of conditions. The large size of this species itself is an expression of abundant nutrient availability.
All the species of diatoms forming the community in each site were salt tolerant and indicators of eutrophic conditions and are capable of withstanding any salt and nutrient enrichment caused by the discharge even during the summer-autumn period. The diatom communities showed no adverse impact in winter, summer and autumn. According to Dr J. John (pers. comm.) the current diatom community should now be considered the natural community of these locations.
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Figure 4-25 Most common dominant diatom species in the system during initial sampling
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Figure 4-26 Profiles of diatom communities immediately above and below discharge sites: winter and summer 2016 and autumn 2017
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Figure 4-27 Scanning electron microscope (SEM) of Cyclotella meneghiniana from the dry-resample 2018 (credit: Dr John Tibby)
45 40 35
30 25 20 15 Abundance (%) Abundance 10 5 0 Burlong Pool US 1.8 DS 3.0 Katrine Pool Site
Cyclostephanos sp Cyclotella meneghiniana Cyclotella pseudostelligera Cyclotella sp. Halamphora veneta All other species
Figure 4-28 Diatom abundances across sites during the summer (low flow) re-sample 2018
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60
50
40
30
20 Sensitivity value (SV) Sensitivity 10
0 US 1.8 DS 3.0 Katrine Pool Burlong Pool Site
Figure 4-29 Diatom sensitivity scores across sites during the summer (low flow) re-sample 2018 only. Error bars represent Standard Deviation.
4.5.3 FISH, MACRO-CRUSTACEANS AND TURTLES Fyke net and box-trap methods were used at four sites. A full breakdown of catches is provided in Appendix A. Fyke nets were set for approximately four hours and yielded fish and shrimp species in all locations. Box trap fish and shrimp counts were excluded from analysis due to low numbers and their inefficiency as a sampling method compared with fyke nets, and because they were only deployed for one season as a result. The counts from both the upstream and downstream fykes for each site were combined for analysis.
The fish and shrimp species recorded in both recessional (November 2016) and low-flow (April 2017) sampling, in order of abundance, were Eastern Gambusia, Gambusia holbrooki (exotic species); Western Hardyhead, Leptatherina wallacei; Glass Shrimp, Palaemonetes australis; and Swan River Goby, Pseudogobius olorum. Additionally, Western Minnow, Galaxias occidentalis was recorded in November. Also in November, two South-western Snake-necked Turtles Chelodina (Macrodiremys) colliei were recorded in fyke nets at Burlong Pool. This species is endemic to southwestern WA and is considered Near Threatened on the International Union for Conservation of Nature (IUCN) Red List for Threatened Species (van Dijk et al. 2012). Both specimens were adult females, measuring: CL = 200 mm, CW = 128 mm, WT = 910g; and CL = 218 mm, CW = 146 mm, WT = 1160 g (CL = carapace length; CW = carapace width; WT = weight).
It is worth noting the catch for the US fyke net at Katrine Pool in April 2017 was exceptionally large, and necessitated subsampling for processing. The haul was sub-sampled by tipping approximately 4/5 of fish in the sampling bucket back into the river. Different species occupy different positions in the water column and were not sub-sampled to the same degree. For example, Pseudogobius olorum is a benthic species and is nearly always found at the bottom of the sampling bucket; therefore, the catch was not extrapolated for this species as the majority remained in the bucket during sub-sampling.
During the low-flow resample (February 2018) large numbers (orders of magnitude higher than previous sampling seasons – see Figure 4-30 and Table 4-10) of several species were caught requiring substantial subsampling. These were, in order of abundance, Eastern Gambusia, Gambusia holbrooki (exotic species); glass shrimp, Palaemonetes australis; Western Hardyhead, Leptatherina Wallacei; and Swan River Goby, Pseudogobius olorum. Two South-western Snake-necked Turtles Chelodina (Macrodiremys) colliei were also recorded in fyke nets at Katrine Pool (Figure 4-31), these were unmeasured, and released immediately after identification.
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Recessional (Nov 2016)
100000
10000
1000 Gambusia holbrooki Leptatherina wallacei 100 Palaemonetes australis
Fish abundance 10 Pseudogobius olorum Galaxias occidentalis 1 Burlong US 1.8 DS 3.0 Katrine Sites
Low (April 2017)
100000
10000
1000 Gambusia holbrooki 100 Leptatherina wallacei Palaemonetes australis Fish abundance 10 Pseudogobius olorum 1 Burlong US 1.8 DS 3.0 Katrine Sites
Low flow re-sample (late February 2018) 100000
10000
1000 Gambusia holbrooki 100 Leptatherina wallacei Palaemonetes australis Fish abudnance 10 Pseudogobius olorum
1 Burlong US 1.8 DS 3.0 Katrine Pool Pool Sites
Figure 4-30 Fish and shrimp species abundance by site and sampling period. Note the log scale for fish abundance.
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Figure 4-31 South-western snake necked turtle caught at Katrine Pool
During the initial survey, the low-flow (April) sampling recorded an overall abundance nearly four-times the recessional period sampling, suggesting a seasonal affect (with decreasing flows, species will often retract into refugia habitat such as that found in pools, which in turn increases catchability of species during this period). Katrine Pool recorded the largest fish counts in both initial sampling seasons; however it is also the largest of the pools sampled with the widest variety of habitats. It was found to be comparable overall to fish counts at Burlong Pool for the recessional sampling period. Burlong Pool was visibly affected by the February 2017 floods, with noticeably reduced macroalgae communities. This could explain in part, the low fish counts observed in April 2017 during low flow surveys, when compared with recessional sampling. During the low-flow resample in February 2018, Katrine Pool again had highest overall abundance, however, this was due to the number of shrimp caught during this sampling period (approx. 14,648). When only teleost fish are counted, Burlong Pool had the highest overall fish abundance (approx. 17,104 at Burlong Pool compared to approx. 9,297 for Katrine Pool) (see Figure 4-32).
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Table 4-10 Species totals at each site across all surveys
Species Burlong Pool US 1.8 DS 3.0 Katrine Pool
Nov Apr Feb Nov Apr Feb Nov Apr Feb Nov Apr Feb ‘18 ‘18 ‘18 ‘18
Gambusia holbrooki 155 50 14304 40 392 7544 161 133 785 200 900 6792
Leptatherina wallacei 187 22 2784 11 464 4592 118 6 641 61 1114 2488
Palaemonetes australis 14 11 2336 47 226 7000 4 1 4 136 607 14648
Pseudogobius olorum 12 5 16 23 15 88 6 1 0 59 157 17
Chelodina 2 ------2 (Macrodiremys) colliei
Galaxias occidentalis - - - 1 - - 1 - - - - -
Total 370 88 19440 122 1097 19224 290 141 1430 456 2778 23947
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12000
10000
8000
6000 Gambusia holbrooki Leptatherina wallacei 4000 Palaemonetes australis 2000 Pseudogobius olorum
Catch estimates of (no. individuals) 0 DS US DS US DS US DS US Burlong Burlong US 1.8 US 1.8 DS 3.0 DS 3.0 Katrine Katrine Pool Pool Pool Pool Site and fyke placement
Figure 4-32 Fish and shrimp species abundance by site and fyke placement – Low flow resample (2018) only Overall there was no indication of impacts to fish abundance or diversity as a result of discharge from the WWTP. Although DS 3.0 recorded much lower numbers of fish compared to upstream sites during the low flow period, this pattern was absent during the recessional sampling suggesting it may be a seasonal effect as opposed to impact from the WWTP. Continuing further downstream, Katrine Pool recorded consistently higher numbers of species than the other sites (including those upstream of the WWTP). The large numbers observed in Katrine Pool when compared to DS 3.0 during the low flow periods may be indicative that it offers more suitable refugia habitat for fish species than DS 3.0 and further investigation would be required to determine the drivers behind these differences.
All fish species recorded are considered to be salt-tolerant and have adapted to a salinity-affected system. The two species most commonly recorded fish species were G. holbrooki and L. wallacei (31,456 and 12,488 total, respectively). Gambusia holbrooki are a highly successful but exotic species widespread through much of Australia. It is known to be extremely tolerant of a wide range of environmental conditions, including salinities ranging from freshwater to fully marine. Leptatherina wallacei are usually only found in brackish estuaries but penetrates far inland in salt-affected rivers (Morgan et al. 2011). Interestingly during the low flow re-sample, the majority of L. wallacei were recorded in the downstream facing fyke nets, indicating upstream movement. The only site where similar numbers were recorded in both upstream and downstream facing fykes was the US 1.8 site. It is worth noting this site is below both the Northam Weir and the confluence with Mortlock River. The weir would be acting as a potential physical barrier during this time, and the Mortlock River is more saline than the Avon River (Mortlock River: 46,967 uS/cm; Katrine Pool: 8,965 uS/cm), which may act as a chemical barrier, forcing movement back downstream within this reach of the river. Pseudogobius olorum, the next most-abundant fish species recorded (399) can tolerate extreme temperature and salinity, and is found in estuaries, rivers, streams, and both fresh and hypersaline lakes.
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5. DISCUSSION
5.1 ENVIRONMENTAL IMPACTS The Northam WWTP has been discharging treated wastewater (TWW) into the Avon River reach at Northam for decades and any impact should be readily apparent from assessing the relative ecology of upstream and downstream sites. Winter is considered ideal for such discharges given the increased dilution factors in the Avon River. However, this work shows that discharge can be carried out on other seasons without significant detriment to the system.
Water balance modelling indicated that the TWW comprises a median value of 0.22% of the Avon River flow though this can increase markedly to a 30-day average of up to 25% during extended periods of TWW discharge during no-flow periods in the river. However, long-term monitoring of nutrients, and the sampling undertaken for the current study, do not show a marked increase in concentrations downstream of the discharge point.
All diatom, fish, and the vast majority of macroinvertebrate species recorded for all sites and surveys are indicative of a saline-affected and impacted system. This applies to sites both above and below the TWW discharge and is consistent with the water quality findings and the multiple impacts to the Avon River. Aquatic ecology surveys do not indicate that the WWTP discharge is having deleterious effects downstream, and in some circumstances additional flow may be beneficial, through dilution, to these communities. However, as noted, under typical operation the effect of TWW discharge is negligible on flow conditions in the Avon River and as such any benefit would be very localised to the discharge point and rapidly diluted by the saline volume of the river itself.
Diatoms are sensitive to ambient changes in water quality, specifically to salinity and nutrient concentrations. The impact of treated wastewater discharge has been successfully assessed by using diatoms (John 2004). Using a “before and after discharge” design, the diatom communities can be compared and any impact from waste water discharge can be determined. The current surveys in the Avon River found diatom assemblages that were well adapted to high salinity and eutrophic conditions, at all sites. The diatom communities showed no adverse impact in high-flow, recessional or low-flow conditions, as a result of WWTP discharge. All the species of diatoms forming the community at each site were salt tolerant and indicators of eutrophic conditions and appear capable of withstanding the incremental nutrient enrichment caused by the discharge even during the summer-autumn period.
Based on diatom assemblages, the recessional period is the only season the system shows some easing of the eutrophic and hypersaline condition of the Avon River. The presence of macro-algae vegetation at all sites assists with the absorption of nutrients and aids to prevent further eutrophication during dry conditions. The natural ephemeral nature of the stream is in some ways conducive to treated wastewater discharge.
Diversity of fish in the streams of southwest WA is generally low compared with the eastern states, however, all fish species recorded are considered to be salt-tolerant and have adapted to a salinity-affected
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system. All are capable of living in brackish conditions or saltier. Surveys indicated that fish species counts and diversity were not impacted downstream of the WWTP.
Similar to the diatom results, the macroinvertebrate communities at all sites and surveys were dominated by salt and drought tolerant species, with no sensitive species recorded, and low in PET taxa. These findings are indicative of disturbed conditions both upstream and downstream of the WWTP discharge, however, the macroinvertebrate fauna of south-west Australian streams is known to be depauperate compared to eastern Australia which includes representatives of the PET groups (Halse et al. 2007).
The Avon River displays environmental impacts from multiple sources. Increased salinity and nutrient enrichment can be associated with the grazing and cropping activities in the catchment, and the aquatic assemblages of the system, both upstream and downstream of the WWTP discharge, are representative of these conditions.
5.1.1 CLIMATE CHANGE The climate change predictions for the south-west of Western Australia, including the Avon Catchment indicate that the annual rainfall will decrease over the next 20-50 years though summer rainfall may increase and rainfall intensity may increase with fewer higher intensity storms (Charles et al. 2010). While this may reduce flows in the Avon River during winter, it will also be likely to increase the demand for water re-use by the Northam council and others at the same time. The nature of the TWW discharge is that it is most active during the wettest period of the year when re-use for Public Open Space (POS) irrigation is not required. Climate change will therefore be likely to have a negative impact on the river health in general though not necessarily due to additional (relatively) inflows of TWW. The feedback between rainfall, groundwater recharge, dryland salinity, salt flushing, and saline groundwater discharge to the river and river health at various periods of the year is complex. Much of the system is controlled by the antecedent conditions, the timing of large rainfall events and the location of local storms within the greater catchment. For example, while the eastern half of the Avon Catchment has significant area, it provides almost no inflows to the Avon River due to the loss to storage in the Yenyenning Lake system (Hennig and Kelsey 2015).
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6. ENVIRONMENTAL IMPACT ASSESSMENT
For calculations of loadings to the Avon River for various changes in the discharge rate, the real recorded daily discharge records for the 2016/17 study period to have been scaled up by the appropriate percent increase. This would equate to a 25% increase to the daily TWW discharge record for an increase from 1.6 to 2.0 ML/d. This allows for temporal variability to be taken into account and represents a more real scenario than a blanket 2.0 ML/d discharge rate applied over the whole assessment period.
It is important to note that the “discharge” or “outflow” volumes mentioned in this report refer to actual measured or planned inflow volumes to the WWTP. There is likely to be some loss to evaporation and infiltration between the inflow and discharge point to the Avon River. As such, the discharge volumes are conservative (over stated).
6.1 SCENARIO 1 – CURRENT PRACTICE AT 1.6 ML/D INFLOW
6.1.1 NUTRIENT LOADING 6.1.1.1 TOTAL NITROGEN SCENARIOS The current loading of TN to the Avon River from the TWW discharge is over an order of magnitude lower than that of the Avon River above the discharge point (Figure 6-1). As outlined in Section 4.4.2, the discharge is typically 0.22% of the Avon River flow up to a maximum of 64% (instantaneous), or 25% on a 30-day average basis, of the flow. These large dilution factors generally mitigate any increase in TN concentration, as well as the importance of any lowering in the TN concentration in the discharge (Figure 6-1). The various scenarios for the reduction in TN from the current 35 mg/L to 22 mg/L or 1.2 mg/L would reduce the influence by up to two or three orders of magnitude less than the upper Avon. The median discharge value could be reduced from the current loading of 16.57 kg/d TN to 0.60 kg/d TN. This however is relative to the median value of 192.3 kg/d TN in the Avon River upstream of the TWW discharge point.
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Figure 6-1 Total nitrogen loading to the Avon River at various discharge quality scenarios for a 1.6 ML/d WWTP. Note the log scale on the y-axis.
6.1.1.2 TOTAL PHOSPHORUS SCENARIOS The Avon River above the TWW discharge has been calculated to have a median loading of 9 kg/d of TP (for the January 2016 to December 2017 period). This is two orders of magnitude higher than the existing TWW discharge at a median value of 0.36 kg/d (Figure 6-2). The existing TWW discharge has a measured median value of 0.72 mg/L and therefore the calculation at a concentration of 1mg/L was similar but increased to a median of 0.50 kg/d TP. Under a scenario where TWW meets the ANZECC guideline of 0.065 mg/L the median value drops considerably to 0.03 kg/d TP.
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Figure 6-2 Total phosphorus loading to the Avon River at various discharge quality scenarios for a 1.6 ML/d WWTP. Note the log scale on the y-axis.
6.1.2 IMPACTS TO THE AVON RIVER 6.1.2.1 WATER QUALITY The existing contribution of the TWW discharge to the Avon River is on the order of 0.22% by volume on a median basis. Long-term monitoring and the sampling undertaken specifically for the current study have shown little if any increase in nutrients or other TWW related parameters downstream in the Avon River.
The potential to reduce Total Nitrogen concentrations in the TWW to 1.2 mg/L would have the effect of reducing the TN inputs to the Avon River by an order of magnitude over current conditions. However, studies of TN loading in relation to aquatic ecosystem health in south-west WA rivers has shown little
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correlation, indicating that other factors are likely to be controlling these systems (DoW 2011a). There are existing cases of algal blooms, including potentially toxic cyanobacteria blooms, in the study area on a regular basis (most summer periods). This would indicate that a reduction in nutrients would be beneficial, but with large catchment scale inputs there may be more cost-effective mechanisms than upgrading the WWTP.
The scenarios investigated for TP represented either an increase (1 mg/L) or decrease (0.065 mg/L). The existing TWW discharge has a measured median value of 0.72 mg/L and therefore the calculation at a concentration of 1mg/L was similar but increased to a median of 0.50 kg/d TP. Under a scenario where TWW meets the ANZECC guideline of 0.065 mg/L the median value drops considerably to 0.03 kg/d TP. This would reduce the loading of TP from TWW to approximately 0.36% from the existing loading of approximately 3.96%. However, given the large capacity of the Avon River to dilute the TWW inflow and the relatively minor contribution to the TP load that the TWW has been calculated to provide, it is considered likely that the current impacts are relatively minor. There may be periods where the TWW discharge increases up to 25% of the Avon River volume in the vicinity of the discharge point, however long-term sampling has shown this to be largely diluted and dissipated within the first 3 km downstream.
6.1.2.2 AQUATIC ECOLOGY The aquatic ecology sampling undertaken for the current study has not indicated a significant impact to the abundance or diversity of diatoms, macroinvertebrates, crustaceans and/or fish downstream of the TWW discharge as far as Katrine Pool (see Section 4.5). It is likely that the salinity of the Avon River and poor water quality in major tributaries such as the Mortlock River are controlling the aquatic ecosystem health of the study reach. While any decrease in TN and/or TP would be likely to be marginally beneficial to the river ecology, the major contributions from the catchment render any benefits relatively minor.
6.2 SCENARIO 2 – INCREASED INFLOW TO 2.0 ML/D
6.2.1 DISCHARGE NUTRIENT CONCENTRATION OPTIONS 6.2.1.1 TOTAL NITROGEN SCENARIOS As with the 1.6 ML/d scenario, the increase to 2.0 ML/d does not change the order of magnitude of TWW contributions to TN loading in the Avon River. The existing TWW discharge represents a median of approximately 8.6% of the loading in the Avon River upstream of the discharge point. This would increase to greater than 11% under a 2 ML/d and 35 mg/L TN scenario. Reducing TN concentrations to 1.2 or 22 mg/L would reduce this to a median of 0.75 kg/d and 13.8 kg/d respectively (Figure 6-3). This equates to a reduction in percentage of total loading to 0.39% (1.2 mg/L TN) and 7.18% (22 mg/L TN). As stated previously, the significant impact of TN on aquatic ecosystem health in south-west rivers has not been established (DoW 2011a) though a general reduction in nutrients may be beneficial in this instance.
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Figure 6-3 Total nitrogen loading to the Avon River at various discharge quality scenarios for a 2.0 ML/d WWTP. Note the log scale on the y-axis.
6.2.1.2 TOTAL PHOSPHORUS SCENARIOS As with Total Nitrogen, the increase of TWW discharge to 2 ML/d will only have a marginal influence on TP loading in relation to the existing loading within the Avon River upstream of the discharge point. The increase in TWW discharge to 2 ML/d would increase the percentage contribution of TP from the TWW to the Avon River from 3.78% under existing conditions to 6.97% at 1 mg/L TP. If the ANZECC guideline target was met however, the TWW discharge loading would decrease to 0.45% at 0.065 mg/L TP.
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Figure 6-4 Total phosphorus loading to the Avon River at various discharge quality scenarios for a 2.0 ML/d WWTP. Note the log scale on the y-axis.
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It is suggested that diatom monitoring occurs in late summer at refugia pools on a three-yearly basis to assess potential changes due to flow variability/climate change and potentially TWW discharge quality. As diatoms are a sessile benthic organism they can show gradients in species assemblages in response to point source changes in water quality such as nutrients and fresh(er) water.
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8. REFERENCES
ANZECC/ARMCANZ (2000). Australian and New Zealand Guidelines for Fresh and Marine Water Quality. Volume 1. Australia and New Zealand.
ARMA/WRC (1999). Avon River Management Programme – WRM11. Avon River Management Authority/Water and Rivers Commission.
Beatty SJ, Morgan DL, Klunzinger M & Lymbery AJ (2010). Aquatic macrofauna of Ellen Brook and the Brockman River: Fresh water refuges in a salinized catchment. Report to Ellen Brockman Integrated Catchment Group.
Blinn DW, Hasle SA, Pinder AM, Shield RJ & McRae JM. (2004). Diatom and micro-invertebrate communities and environmental determinants in the western Australian wheatbelt: a response to salinization. Hydrobiologia, 528, 229-248. doi: 10.1007/s10750-004-2350-8
Canedo-Arguelles M, Kefford BJ, Piscart C, Prat N, Schafer RB and Schulz CJ (2013). Salinisation of rivers: An urgent ecological issue. Environmental Pollution 173 (2013) 157-167.
Charles SP, Silberstein R, Teng J, Fu G, Hodgson G, Gabrovsek C, Crute J, Chiew FHS, Smith IN, Kirono DGC, Bathols JM, Li LT, Yang A, Donohue RJ, Marvanek SP, McVicar TR, Van Niel TG and Cai W. (2010). Climate analyses for south-west Western Australia. A report to the Australian Government from the CSIRO South-West Western Australia Sustainable Yields Project. CSIRO, Australia. 83 pp.
Chessman B. (2003). SIGNAL 2 – A scoring System for Macro-invertebrates (‘Water Bugs’) in Australian Rivers. Monitoring River Health Initiative Technical Report no 31. Commonwealth of Australia, Canberra.
Chessman B, Bate N, Gell PA & Newall P. (2007). A diatom species index for bioassessment of Australian rivers. Marine and Freshwater Research, 58, 542-557. doi: 10.1071/MF06220
Chubb CF, Hutchins JB, Lenanton RCJ & Potter IC (1979). An Annotated Checklist of the Fishes of the Swan-Avon River System, Western Australia. Rec. West. Aust. Mus., 1979, 8(1).
Conrick D, & Cockayne B (2001). Queensland Australian River Assessment System (AusRivAS) Sampling and Processing Manual. Brisbane: QLD Department of Natural Resources and Mines.
DBCA (2017). Priority Ecological Communities for Western Australia - Version 27. Species and Communities Branch, Department of Biodiversity, Conservation and Attractions, Government of Western Australia. 30 June 2017.
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DEC (2012). Aquatic invertebrate communities in Gwambygine Pool following dredging in 2010. Report to Wheatbelt Natural Resources Management Inc. by Melita Pennifold, Science Division, Department of Environment and Conservation, Western Australia. April 2012.
DoW (2007). Assessment of the status of river pools in the Avon catchment. Report No. WRM 47, December 2007. Department of Water, Government of Western Australia.
DoW (2008). Priority tributaries of the Avon River basin: a process to prioritise tributaries for condition assessment. Volume 1 Avon and Mortlock Catchments. Water resource management series, Report no. WRM 51. Department of Water: Perth, Western Australia.
DoW (2011). Assessment of the ecological impacts of sediment removal at Gwambygine Pool, Katrine Pool and Reserve Pool. Unpublished report to Wheatbelt NRM. Department of Water, September 2011.
DoW (2011a). The Framework for the Assessment of River and Wetland Health (FARWH) for flowing rivers of south-west Western Australia – Method Development. Water Science Technical Series, Report No. WST 40. September 2011.
DoW (2011b). The Framework for the Assessment of River and Wetland Health (FARWH) for flowing rivers of south-west Western Australia – Project summary and results. Water Science Technical Series, Report No. WST 39. September 2011.
Fore L.S. and Grafe C. (2002). Using diatoms to assess the biological condition of large rivers in Idaho (USA). Freshwater Biology 47: 2015-2037.
GHD (2009). Avon Catchment Council - Report for Surface Water Management and Self Sufficiency - Environmental Water Flow Requirements. June 2009.
Halse S, Scanlon M, Cocking J, Smith M, & Kay W (2007). Factors affecting river health and its assessment over broad geographic ranges: The Western Australian experience. Environmental Monitoring and Assessment, 134, pp. 161-175.
Hennig, K. and Kelsey, P. (2015). Avon Basin hydrological and nutrient modelling. Water Science Technical Series, report no. 74. Department of Water, Government of Western Australia.
John J (2000). A guide to diatoms as indicators of urban stream health. Land and Water Resources Research and Development Corporation.
John J, Buston F, & O’Connor J (2002). Diatom Biomonitoring of Wastewater Discharge on Streams at Katanning, Kojonup and Narrogin. Wetland Research Group Report 1. Curtin University.
John J (2004). Diatom assemblages as indicators of wastewater discharge in a temporary stream in Western Australia. In Seventeenth International Diatom Symposium 2002, Aug 25, 2002. Ottawa, Canada: Biopress Ltd.
John J (2012). A Diatom Prediction Model and Classification for Urban Streams from Perth, Western Australia. Germany: Koeltz Scientific Books.
Morgan D, Beatty S, Klunzinger M, Allen M & Burnham Q (2011). A Field Guide to Freshwater fishes, Crayfishes and Mussels of South-western Australia. SERCUL, Beckenham, WA. Negus P, Steward A & Blessing J (2013). Queensland draft macroinvertebrate guidelines: Murray–Darling and Bulloo catchments, November 2013 – Draft for Comment. Department of Science, Information Technology, Innovation and the Arts, Queensland Government: Brisbane, QLD
Oeding S, & Taffs KH (2015). Are Diatoms a Reliable and Valuable Bio-Indicator to Assess Sub-Tropical River Ecosystem Health? Hydrobiologia, 758 (1), 151–69. doi:10.1007/s10750-015-2287-0.
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Oren, A. and Vlodavsky, L. (1985). Survival of Escherichia coli and Vibrio harveyi in Dead Sea water. FEMS Microbiology Ecology 31 (1985) 365-371.
Pennifold M (2007). Avon River Pool Invertebrate Survey. Report to the Species and Communities Branch, Department of Environment and Conservation. Science Division, Department of Environment and Conservation, Western Australia – 7/2007.
Pinder A (2009). Aquatic Invertebrate Communities in Avon and Dale River Pools. Report to Species and Communities Branch, Department of Environment and Conservation.
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Turtle Taxonomy Working Group [van Dijk, P.P., Iverson, J.B., Shaffer, H.B., Bour, R., and Rhodin, A.G.J.] (2012). Turtles of the world, 2012 update: annotated checklist of taxonomy, synonymy, distribution, and conservation status. Chelonian Research Monographs No. 5, pp. 000.243–000.328, doi:10.3854/crm.5.000.checklist.v5.2012.
WALGA (2015). Shire of Northam - Local Biodiversity Strategy. February 2015. van Looij E (2009). WA AusRivAS sampling and processing manual. Water Science Technical Series Report No. 13, Department of Water, Western Australia.
WA Government (1963). Salinity Problems in Western Australian Catchments with particular reference to Wellington Dam – compiled by W H Power. File PWWS 251/51, Appx 9.
Water Corporation (2017). Annual Environmental Report. Northam WWTP. Reporting date: 1 July 2016 to 30 June 2017. Document prepared for Department of Water and Environmental Regulation. Water Corporation: Leederville, Western Australia.
Wheatbelt NRM (2013). Avon River – Strategic Review. https://www.wheatbeltnrm.org.au/sites/default/files/basic_page/files/Avon%20River.pdf
Wood WE (1924). Increase of Salt in Soil and Streams following the Destruction of Native Vegetation. J. Roy. Soc. WA 10 (7): 35-47.
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APPENDIX A. AQUATIC ECOLOGY CATCH DATA
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APPENDIX B. WATER QUALITY DATA
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Table Apx B-1 Field Parameters
Table Apx B-2 Major ions
Table Apx B-3 Dissolved metals
Table Apx B-4 Total metals
Table Apx B-5 Nutrients
Table Apx B-6 Total ions, chlorophyll and hydrocarbons
Table Apx B-7 Microbiology and surfactants
Table Apx B-8 Herbicides and Pesticides
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