The Influence of Concrete and Urban Water Quality On The Growth of An Invasive Weed (Salix Spp.) in One of Australia’s Most Endangered Aquatic Environments.


 Riparian vegetation along urban streams and wetlands is frequently dominated by invasive weeds. Elevated nitrogen and phosphorous in urban waters and soils are well-known to encourage invasive urban weeds, but this research demonstrates that other urban geochemical contaminants may also be influential. Previous studies have demonstrated that the dissolution of urban concrete is a poorly recognised source of modified water and soil geochemistry, which may enhance the growth of some invasive weeds. This study investigated the relationship between urban water quality and the growth of an invasive urban riparian weed, willow (Salix spp.) to examine the contribution of concrete materials. The study used water from a wetland in the Greater Blue Mountains World Heritage Area. These wetlands have a unique biodiversity but are fragile and susceptible to degradation from human activity. Many are in urban catchments and are frequently dominated by invasive weeds, including Salix spp. In this study, willow cuttings were grown in a laboratory using four water treatments: pristine, urban, and pristine water exposed to two different concrete materials. The urban and concrete water treatments had higher pH, salinity, calcium, potassium, and higher concentration of several metals and were associated with increased growth of Salix spp. We suggest that the modification of urban water and riparian soil chemistry by urban concrete materials may contribute to the success of invasive species in urban wetlands and riparian zones. Some metals (barium, strontium) were present in urban water and in pristine water exposed to concrete and bioaccumulated in plant tissue.


Introduction
Riparian zones alongside streams in urbanised areas frequently exhibit modi ed vegetation communities when compared to non-urbanised areas ( Riparian vegetation plays a key role in stabilising, binding and reinforcing banks (Gurnell, Lee and Souch 2007), as well as regulating stream temperature, trapping and ltering nutrients and pollutants, and contributing organic matter to streams (Wenger et al. 2009). Thus, changes in riparian vegetation communities can have cascading ecosystem effects and can result in erosion, unstable channels, and increased in ux of pollutants to the waterway (Gurnell, Lee and Souch 2007). Dominance of invasive plant species in riparian vegetation can also have detrimental impacts through reduction of habitat value for endemic biota and particularly for endangered fauna species (Samways and Taylor 2004).
Many factors likely contribute to the success of non-native species in urban waterways. Exotic species are generally able to invade urban stream ecosystems due to a series of hydrological, physical, and chemical changes associated with USS (Paul and Meyer 2001). An important factor is proximity to urban areas, which allows propagules to enter the stream from nearby human habitation (King and Buckney 2000). Urban areas also provide numerous convenient transport paths for weed propagules, including waterways, vehicles, humans, and domestic animals (Akasaka, Osawa and Ikegami 2015).
Thus, exotic species present in populated centres potentially have several mechanisms available to spread to nearby ecosystems. However, the success of exotic plant species in urban riparian zones is also strongly in uenced by the chemical differences observed in water and soil between urban and non-urban areas (Leishman and Thomson 2005).
Urban waterways are generally chemically degraded compared to waterways in pristine non-urban areas ( Similarly, urban soils exhibit a variety of chemical differences when compared to soil in more pristine areas. Multiple studies have noted that areas with higher proportions of invasive species had soil that was higher in pH and salinity as well as highly enriched with potassium, phosphorous, and calcium (Grella, Renshaw and Wright 2018;Park et al. 2010; Shaw and Reeve 2008). Nutrients such as nitrogen and phosphorous are linked with increased weed growth and higher concentrations of both are associated with riparian weed invasions (Leishman, Hughes and Gore 2004; Leishman and Thomson 2005). Modi ed urban riparian soil chemistry may inhibit native species that are less tolerant to these conditions and encourage the establishment of exotic species that are more adapted to these urban environments (King and Buckney 2000). This idea is supported by Leishman and Thomson (2005), who observed that native plant seedlings were killed by elevated soil phosphorus concentrations.
Urban water and soils also commonly contain a wide range of pollutants including pesticides and heavy metals (Paul and Meyer 2001). This has been termed urban geochemistry and common metals recorded in urban streams at elevated levels include cadmium, chromium, copper, lead, and zinc (Wenger et al. 2009). Depending on their bioavailability, heavy metals can transfer further into the ecosystem and can have a variety of effects on plant growth and processes (Ugulu 2015).
While many heavy metals are important in plant and animal biochemical and physiological functions, they can also be toxic at higher concentrations (Tchounwou et al. 2012). For example, chromium toxicity in plants has been reported (Oliveira 2012), and copper can be toxic to aquatic invertebrates (Tchounwou et al. 2012). Heavy metals can also affect a variety of plant processes, including tissue development, germination, and photosynthesis (Oliveira 2012).
There is growing recognition that urban geochemistry can be associated with weathering of urban concrete materials (Kaushal et al. 2013;Tippler, Wright and Hanlon 2012). Recirculation of non-urban water through pipes composed of concrete materials has shown that exposing water to typical urban concrete materials can result in a rapid increase in water pH, salinity, calcium, potassium, and bicarbonate (Davies et al. 2010;Grella et al. 2014). Recycled concrete aggregates (RCAs) are a widely used recycled building material that are often used in urban construction and drainage settings. When RCAs have been exposed to water in laboratory settings, experimental results have documented rapidly increasing pH and the release of a suite of ionic and metallic elements, sometimes rapidly reaching ecologically hazardous concentrations ( Previous studies have reported that water quality in naturally vegetated non-urban BMUS is acidic (mean pH<5) and dilute (<41 µS/cm), with a very low buffering capacity and undetectable concentrations of calcium and magnesium (Belmer, Wright and Tippler 2015; Carroll, Reynolds and Wright 2020). In contrast, urban BMUS have higher pH, EC, and increased concentrations of all major ions (Belmer, Wright and Tippler 2015; Carroll, Reynolds and Wright 2020). Carroll, Reynolds and Wright (2020) investigated the concentration of 17 metals in urban and non-urban BMUS and found that two metals (barium and strontium) were present at much higher concentrations in water and sediment of urban BMUS. The difference between urban and non-urban BMUS was largest for strontium. It was measured in water in urban BMUS (mean 45 µg/L), which was more than 30 times greater than non-urban BMUS (mean 1.4 µg/L) (Carroll, Reynolds and Wright 2020). Both Carroll, Reynolds and Wright (2020) and Belmer, Wright and Tippler (2015) supported the hypothesis that the geochemical pattern of contamination in urban BMUS appears to be in uenced by dissolution of urban concrete materials (Davies et  Given that concrete is becoming more widely recognised as a source of geochemical modi cation in urban waterways (Davies et al. 2010;Grella et al. 2014;Wright et al. 2011), it is important to investigate how these changes in water and soil chemistry in urban waterways are likely to affect riparian vegetation. Urban concrete infrastructure is being increasingly recognised as a source of metal contamination (Purdy and Wright 2019; Rahman et al. 2014) and this has many implications for ecosystems, including potentially affecting plant processes by contributing metals to urban water and soils (Ugulu 2015). Plants accumulate pollutants in their tissues and plants in urban areas near high tra c streets and residential areas frequently have higher concentrations of metals in their tissues (Buszewski et al. 2000). Metals are also frequently present in urban streams (Meyer, Paul and Taulbee 2005;Wenger et al. 2009) and the effects of having such metals and modi ed ionic composition in an urban riparian ecosystem is poorly understood. Further investigation is required to determine whether heavy metals present in stream water become bioavailable and are taken up into plant tissue, and thus whether they may in uence riparian vegetation. Additionally, it is hypothesised that, if heavy metals are bioavailable, they may travel further into the food chain and ecosystem and could be used as tracers or markers to measure the level of urban concrete exposure in ecosystems.
This study aimed to: 1. Explore how urban water and pristine water exposed to concrete in uence the growth and development of a known invasive riparian weed species in the Blue Mountains region, Salix spp.
2. Investigate whether metals leached into water via concrete exposure are taken-up by Salix spp. into its tissue and whether they can be used as metal 'tracers' for concrete exposure.

Water treatment chemical differences
Water chemistry differed between the four different water treatments at the outset of this study (Purdy et al. 2020; Table   1). The four treatments used water obtained from two different BMUS wetlands. One treatment was sourced from surface water from a BMUS in a catchment that was partly occupied by urban land. This was the Bullaburra BMUS, which had been affected by recent modi cations to the local stormwater system, built to drain a nearby local highway (Belmer, Wright and Tippler 2015). This water was the urban treatment.
Water was also collected from the surface water of a reference BMUS in a naturally vegetated area, managed and protected as part of Greater Blue Mountains World Heritage Area (Belmer, Wright and Tippler 2015). This near-undisturbed BMUS, near Mount Hay, had no urban development in its catchment (Belmer, Wright and Tippler 2015). This water was used as the pristine reference treatment, with additional water collected from this location as the source water for exposure to different forms of concrete materials (Purdy, Reynolds and Wright 2020).
Two sub-samples of additional pristine reference water were exposed to different concrete materials according to detailed methods available in an associated study (Purdy, Reynolds and Wright 2020), which produced two of the water treatments used in this current study. Purdy, Reynolds and Wright (2020) produced these water treatments by recirculating pristine reference BMUS water through two different types of concrete materials for 60 minutes. One was an unused steelreinforced concrete water pipe and produced concrete whole pipe water treatment. The other produced RCA water by recirculating pristine reference water through a PVC gutter containing small (<20mm) concrete fragments called recycled concrete aggregate ( The RCA water treatment had the highest mean values for ten of the 13 water chemistry attributes that were measured for all water treatments prior to the growth experiment beginning ( Table 1). The pristine reference water only had detectable results for seven of the 13 water quality attributes ( Table 1). The four water treatments all had different pH and EC values ( Table 1). The pristine reference water had lowest mean pH, at 4.89, while the RCA water had the highest mean pH at 8.91.
Similarly, pristine reference water was the most dilute (mean 27.88 µS/cm), whilst the mean EC of RCA water was more than 11 times higher and was highest of the treatments in this study (340.0 µS/cm).
Differences in the concentration of major cations were observed across the four different treatments prior to the start of the growth experiment (Table 1) (Purdy, Reynolds and Wright 2020). The only cation detected in pristine reference water was sodium (mean 3.7 mg/L). The highest sodium concentration was found in RCA water (mean 21.5 mg/L). Calcium was the dominant cation in RCA water (mean 34.6 mg/L) and was lower in urban water (mean 12.33 mg/L) and concrete pipe water (mean 8.49 mg/L). Potassium was not detected in reference water but was measured in urban water (mean 2.04 mg/L), in concrete pipe water (mean 2.93 mg/L) and was highest in RCA water (13.7 mg/L). Magnesium was not found at detectable concentrations in pristine reference water or concrete pipe water but was in urban water (mean 1.34 mg/L) and RCA water (mean 0.97 mg/L).
Water samples were also tested for seven metals prior to the start of the growth experiment (Table 1). Aluminium was the most abundant metal, with the lowest concentration in urban water (mean 40 µg/L), followed by concrete pipe water (mean 308 µg/L), pristine reference water (mean 370 µg/L) and largest in RCA water (mean 654 µg/L). Two of the metals (manganese and strontium) were not detected in the pristine reference water. Barium was detected in the highest concentration in urban water (mean 16.3 µg/L), then RCA water (14.6 µg/L). Copper was highest in water recirculated through concrete materials (concrete pipe water mean 9.6 µg/L; RCA water mean 12.0 µg/L). The largest differences in metal concentrations were found for strontium, which was not detected in reference water, compared to a mean of 43 µg/L in urban water, 13.9 µg/L in concrete pipe water and highest at 125.6 µg/L in RCA water.

Study species and experimental setup
The invasive willow species (Salix spp.) was chosen as a study species, as it grows quickly and is easy to grow in controlled laboratory conditions. Additionally, willows in the genus Salix were amongst nine others listed as problematic weed species in the recommended declaration of BMUS as a vulnerable ecological community (NSW Scienti c Committee 2007). Salix spp. is also listed as a problematic weed species across urban streams and wetlands in the Blue Mountains and wider Sydney region (Blue Mountains City Council (BMCC) 2016) due to its ability to spread vegetatively and by seed (Webb and Erskine 2003). Willows are also common in urban, rural and other riparian zones across Australia, as they were historically planted for aesthetic bene ts and to control streambank erosion caused by the prior removal of native species (Webb and Erskine 2003). Another reason for using willows as a test species was that previous research had demonstrated their ability to uptake and accumulate heavy metals in their tissue (Meers et al. 2007).
Willows have well-known detrimental effects on many Australian riparian ecosystems, including reduced channel capacity, channel diversion and erosion, increased sediment loads and increased ooding, which can affect water quality and river health (Stokes and Cunningham 2006). Hence, results of this study could help understand how contamination associated with human activity can in uence the growth of this species. This information could have management implications for Salix spp. as it is classi ed as a noxious weed and is listed as a weed of national signi cance (NSW DPIE 2021). NSW DPIE considers species of willow (Salix spp.) as one of the worst weeds in Australia due to their invasive abilities and their economic and environmental impacts to streams and rivers, in addition to their ability to out-compete native species and dominate riparian vegetation (NSW DPIE 2017).
120 willow cuttings were randomly selected with 30 in each of the four water treatments (Figure 1). Branches of Salix spp. were collected from ve parent willow trees located across the Hawkesbury-Nepean and Blue Mountains regions. More than one parent tree was selected to ensure that any growth differences across treatments were not due to the individual characteristics of any one parent tree. The branches were trimmed to approximately 30 cm lengths and were stripped of all leaves and shoots. A pilot study that was conducted in preparation for this experiment revealed that thicker willow cuttings showed improved growth (Purdy and Wright 2019). Thus, for this study, only cuttings with diameters from 0.3-0.6 cm were selected.
The willow cuttings were placed into one of four different water treatments; urban water; reference; concrete pipe water, and RCA water. Each cutting was placed individually into a cleaned and rinsed plastic 500 mL beaker, which was lled with 400 mL of the respective water treatment. The top of the beaker was covered with a plastic wrap lm to minimise evaporation, with a small hole for the cutting. Willow branch collection, preparation of cuttings, and placement in beakers containing treatment solution occurred on a single day (Figure 1).
The plants were randomly assigned positions in a controlled environment growth room, where they were exposed to 24hour light (Philips 'cool white' TLD36W/840) and approximately constant temperature. Minimal temperature uctuations occurred throughout the growth period, from 20-21 degrees Celsius, measured with a Livingstone maximum-minimum thermometer. Beakers were topped up to maintain a volume of 400 mL with deionised water on a fortnightly basis. The plastic lm was reapplied after water testing and topping up water level.

Harvest and plant growth measurements
Plants were harvested after 7.5 weeks of growth. For each plant, maximum shoot and root length, stem height, the length of ve random adult leaves, and whole plant wet weight were measured ( Figure 2). Maximum root length was measured as the length of the longest root, with the root straightened perpendicular to the main stem or trunk of the plant. Maximum shoot length was measured as the length of the longest stem growing from the main willow trunk and leaf length was the distance of the entire blade from base to apex. Stem height refers to the total length of the main trunk of the plant, from base to tip. Wet weight was measured using an electronic balance (AND FX-3200) and separate weigh-boats were used for each plant, to prevent cross contamination.
Leaf area measurements were taken from ve randomly selected plants from each treatment group. Leaves were plucked gently from the plant, placed in a scanner, and scanned in colour (Figure 3). The scanner was wiped with ethanol and allowed to dry in between plants to ensure that no contamination between plant tissue occurred. A ruler was placed on the scanner for scale. Leaf area measurements were processed using ImageJ software (Rasband 1997(Rasband -2018) and the protocol developed was based on Glozer (2008).
All plants were stored in unused paper bags and kept at 20 degrees Celsius for 1.5 weeks, after which they were placed in an incubator (Labec incubator S4218) for 9 days at 38.5 degrees Celsius. Plants were removed when the weights of ve random plants taken a day apart varied by less than 0.05 g. For each plant, whole dry weight was measured. The shoots (including leaves) and roots were then separated from the main trunk and their dry weight was taken, to measure the weight of the new growth of the plant during the experimental period. Shoots and leaves were then weighed separately, as were roots. For all weight measurements, an electronic laboratory balance was used (AND FX-3200).

Analysis of metals in plant tissue
For analysis of metals in the dried plant tissue, each treatment group (pristine, urban, whole concrete, RCA) was divided into ve replicates. Each replicate contained the shoot and root tissue from six plants.
The shoots and leaves of the tissue from six plants in each subgroup were ground up using an electronic coffee grinder (Sunbeam Multigrinder II). Shoots and leaves were placed in the coffee grinder and were ground for two to four minutes in 20 second bursts until a ne powder was produced. Each subgroup was then sieved through an 850-micron sieve and stored in a sterile plastic tube ( Figure 5). In between grinding of samples, the grinder and sieve were thoroughly cleaned and then rinsed with deionised water and dried with a hairdryer, to prevent cross-contamination.
The plant tissue samples were then sent to a NATA accredited laboratory, where concentrations of 16 metals was measured: arsenic, barium, cadmium, chromium, copper, manganese, molybdenum, nickel, lead, titanium, strontium, lithium, aluminium, iron, and zinc. Metal concentrations were determined using ICP-AES (Envirolab Services Pty Ltd, personal communication 11 July 2019).

Statistical analyses
Data was analysed using RStudio (RStudio Team 2016). Normality of data was assessed using the Shapiro-Wilks test and homoscedasticity was evaluated using Levene's test, using the 'car' package (Fox and Weisberg 2019). As data was nonnormal (p < 0.05) and variances were not equal (p < 0.05), Kruskal-Wallis tests were used as a non-parametric alternative to assess whether differences in growth and in tissue metal concentrations were signi cant between treatments. The threshold of signi cance chosen was 0.05, thus any results with p values less than 0.05 were considered signi cant. A Dunn's posthoc test with Bonferroni p-value correction in the 'FSA' package (Ogle, Wheeler and Dinno 2019) was used for any measurements that were shown to signi cantly differ to assess which treatment groups differed. All graphs and gures were produced using Microsoft Excel (Microsoft 2019).
For growth measurements, treatment group averages were calculated based on the 30 plants in each treatment group (n=30, total n=120). For leaf area measurements, averages were calculated based on ve random plants chosen from each treatment group (n=5, total n=20). For the metal tissue content, averages were calculated based on ve samples for each treatment group (n=5, total n=20), which were each composed of 6 random plants of the same treatment group.

Plant growth
After 7.5 weeks in the controlled environment, the leaf, shoot and root growth from the willow cuttings differed signi cantly between treatment groups for all measured growth parameters ( Figure 2; Table 2). For length and dry weight measurements, the plants grown in RCA water were both longest and the heaviest, followed by plants grown in urban BMUS water, then plants grown in whole pipe recirculated water, and, nally, plants grown in pristine BMUS water. The mean maximum length of willow cutting shoots (8.34 cm) and roots (11.79 cm) from pristine water treatment plants were shortest, while the longest were found in plants grown in the RCA treatment (shoots 17.3 cm; roots 27.47 cm; Table 2). The same trend was observed for the dry weight of new willow growth. The mean dry weight of new growth was lightest for pristine treatment plants, comprising shoots and leaves (0.40 g) and roots (0.07 g). In contrast, the mean dry weight of new growth was heaviest for RCA treatment plants, comprising shoots and leaves 0.78 g and roots 0.23 g ( Table 2).
There were also large differences in leaf measurements, which varied signi cantly according to treatment (Table 2; Figure   3). There were signi cant differences between leaf parameters (leaf area and number of leaves) of pristine treatment plants and urban treatment plants. Pristine treatment plants had the lowest mean number of leaves (15.8) contrasting markedly with the number of leaves from plants from the other three treatments, which were from two to three times greater, ranging from a mean of 52.8 leaves (RCA treatment) and 66.6 leaves (urban treatment) ( Table 2; Figure 3). The total leaf area of pristine plant cuttings (mean 475.68 cm 2 ) was the lowest of all the treatments. In comparison, mean leaf area was between 53.9% greater (concrete whole pipe treatment = 732.3 cm 2 ) and 85.1% greater (urban plants = 880.3 cm 2 ).

Metal concentrations in willow tissue
Differences in mean content of seven metals were observed in the tissue of plants grown in the four different water treatments (Table 3; Figure 5). The concentration of three metals (barium, strontium, and aluminium) in plant tissue varied signi cantly according to treatment (Table 3). A Dunn's post-hoc test revealed that barium concentrations differed signi cantly between pristine and RCA treatment plants, while strontium concentrations were signi cantly different for pristine and RCA treatment plants and for RCA and concrete pipe treatment plants.
Barium concentrations recorded the greatest overall percentage increase in metal concentrations in plant tissue grown in different treatments (Table 3). When comparing the barium content in plant tissue grown in pristine BMUS water to plants grown in the other three treatments as a percentage increase, the greatest percentage increase was between pristine (mean barium 16.4 mg/kg) and RCA plants (mean barium 41.6 mg/kg) with an approximate increase of 154% (Table 3). The trend was similar for strontium concentrations between pristine (mean 28.8 mg/kg) and RCA plants (mean 68.4 mg/kg), with an increase of approximately 137% (Table 3). For aluminium, the pattern was the opposite, with the largest concentrations found in pristine plants (mean 11.6 mg/kg) compared to all other treatments (urban, whole pipe and RCA), for which mean aluminium concentrations ranged from 8.0 -8.8 mg/kg.

Discussion
Willow cuttings in this study demonstrated increased growth in treatments of urban water and pristine water exposed to concrete, compared to willows grown in unmodi ed pristine water. Overall, from all measured growth parameters, willows grown in RCA recirculated water showed the greatest growth, followed by plants grown in urban BMUS water, then those grown in whole concrete pipe recirculated water, and, nally, willows grown in pristine BMUS water. Results suggest that exposure of water to concrete materials contribute to favourable water geochemistry for the growth of the willow, Salix spp., which is regarded as an Australian weed of national signi cance. Salix spp. is also identi ed as an invasive weed of concern for the conservation of endangered BMUS ecological communities (NSW Scienti c Committee 2007).
These ndings are supported by an earlier pilot study, which recorded increased growth of willow cuttings in reference BMUS water circulated through a concrete pipe for two hours, when compared with cuttings in reference water that had not been exposed to concrete (Purdy and Wright 2019). To the best of our knowledge, there are no other laboratory studies investigating how water exposed to concrete may affect the growth of an invasive weed species and, thus, the results of this study and our previous study (Purdy and Wright 2019) appear to be unique. These results add to the growing knowledge of how widespread urban materials, such as concrete, can affect the natural environment.
Urban riparian zones and urban wetlands are frequently dominated by vigorous growth of weed species, often to the detriment of native species (Ehrenfeld and Schneider 1991; Grella, Renshaw and Wright 2018; Leishman, Hughes and Gore 2004). Previous research widely attributes the dominance of invasive plant species in urban riparian zones to elevated nitrogen and phosphorous levels in water and riparian soil (Paul and Meyer 2001). However, in the case of BMUS, weed invasions are prevalent in wetlands of urban catchments that do not have increased nitrogen and phosphorous concentrations in the water (Carroll, Wright and Reynolds 2018). The results of this study highlight other anthropogenic factors, such as higher pH and increased concentration of metals and ions, that contribute to favourable growth conditions for urban invasive weeds.
The other key nding in this study is that water exposed to concrete can increase the growth of invasive weeds. This adds further support to the theory that urban concrete materials in the built environment can be a potential source of modi ed geochemistry in urban waters and soils and therefore can contribute to the chemical and the ecological degradation of urban streams and wetlands ( Water in the four treatments used to grow willow cuttings in this study had varying pH levels, from a mean of 4.89 in pristine reference BMUS water to 8.91 in the RCA treated water. Urban BMUS water had a pH of 7.53, while water exposed to a whole concrete pipe had a pH of 7.42. This was in accordance with previous studies of pristine non-urban BMUS water that often record pH <5, with urban BMUS generally reporting pH >6 (Belmer, Wright and Tippler 2015; Carroll, Reynolds and Wright 2020). It is well known that pH plays an important role in uencing plant growth through the bioavailability and uptake of nutrients and toxicity of metals (Alam, Naqvi and Ansari 1999; Charman and Murphy 2007; Murrmann and Peech 1969; Wang et al. 2006). Previous research has reported that pH differences can be highly in uential to plant growth, including weed growth (Buchanan, Hoveland and Harris 1975). For example, an experimental study using tomato seedlings reported that when soil pH of was at both 4 and 8, plant growth was diminished, but at a pH of 6, plant growth was optimal (Kang et al. 2011). Riparian soils have also recorded substantial pH differences in urban versus nonurban catchments (Grella, Renshaw and Wright 2018). Grella, Renshaw and Wright (2018) reported that the mean riparian soil pH was 5.0 in catchments with low coverage of urban development and was 6.49 in catchments with highest coverage of urban development. That study also found that riparian soil pH was the single most highly correlated factor associated with variation in riparian vegetation, with higher pH urban soils having greater coverage by invasive weeds (Grella, Renshaw and Wright 2018). It is likely that pH differences in water treatments in this study contributed to the observed growth differences.
In the current study, elevated potassium and calcium concentrations in urban and both concrete water treatments were also likely contributors to the increased growth of willow cuttings observed in these treatments. Calcium and potassium are both regarded as essential plant nutrients and are integral for numerous functions essential for plant growth (Burstrom 1968;Groffman and Fisk 2011;Schachtman and Schroeder 1994). Potassium is very important for plant growth as it is generally the most abundant cation in plants (Schachtman and Schroeder 1994). Two previous investigations of soils and weeds in sandstone-derived soils in the Sydney area (Grella, Renshaw and Wright 2018; King and Buckney 2000) reported increased weed growth in urban soils with higher calcium and potassium content, compared to naturally forested riparian soils. Another study (Carroll, Reynolds and Wright 2020) tested the major ion contents of plant foliage growing in urban and non-urban BMUS. Potassium accounted for a mean of 3129 mg/kg of dried plant tissue in non-urban BMUS foliage, representing 33% of combined major cations content. The same study found that plant tissue from urban BMUS foliage had substantially (5.5 times more) greater potassium content (mean 17740 mg/kg) than non-urban plants, which accounted for 61% of the combined major cations content. The strontium and barium contents in tissue samples from willow cuttings grown in the three urban/concrete water treatments (urban, concrete pipe and 20 mm RCA) were more elevated than those grown in the reference treatment. This was similar to a previous pilot study for the current research (Purdy and Wright 2019). The previous study also grew willow cuttings for 7 weeks in two water treatments. One was water from a non-urban BMUS and the other treatment was a sample of water from the same BMUS that was recirculated through an unused concrete pipe. The major difference between this study and the previous study was that the water in the previous study was recirculated for 120 minutes, double the 60 minutes used in the current study. The previous study also recorded a large increase in both barium and strontium in willow tissue grown in the concrete pipe treatment. Another study of water quality in urban and non-urban BMUS also measured barium and strontium at much higher concentration (barium 4.5 times and strontium 30 times) in urban compared to non-urban locations (Carroll, Reynolds and Wright 2020). The water quality and bioaccumulation results suggest that barium and strontium may be a useful as a marker of concrete exposure, based on the results of this study and previous research (Carroll, Reynolds and Wright 2020; Purdy, Reynolds and Wright 2020; Purdy and Wright 2019).
Urban concrete materials are likely to contribute to a substantial proportion of strontium and barium in urban waters. Strontium particularly is often one of the most abundant trace elements in many concrete materials (Graham, Goguel and St John 2000; Long, Zhang and Luo 2019) depending on the source and formulation of ingredients used. Strontium is likely to be in many Australian concrete products due to the widespread use of coal-derived y-ash in Australian concrete materials. It has been estimated that Australian concretes commonly contain up to 5% y-ash (Ash Development Association of Australia (ADDA) 2009). Fly-ash itself often has a substantial strontium content, with United States y ash reported to contain an average of 1334 ppm of strontium (Straughan et al. 1981). Trace metal analysis of Australian yash revealed that strontium was the second most abundant metal (360 ppm), second to barium (490 ppm) (Ward et al. 2009). Our suggestion that strontium may be a marker of urban and concrete exposure is not completely novel. It has previously been suggested that strontium can be used as a soil, water, and plant marker of coal combustion residue, particularly from y ash (Hurst, Davis and Elseewi 1991; Straughan et al. 1981). Several studies have revealed that plants are able to bioaccumulate strontium into leaf and root tissue, sometimes damaging the plant (Burger and Lichtscheidl 2019). Straughan et al. (1981) has also reported that strontium can be readily bioavailable, as demonstrated when plants grown in soil containing an 8% y ash mixture accumulated strontium in leaf tissue at up to 300 ppm. That was about 2.5 times greater than the maximum strontium content of 81 ppm recorded in the current study, in the willow leaf tissue grown in the RCA water treatment.
The current research builds on previous studies and highlights the likely contribution of concrete urban materials to the geochemistry of urban waterways (Davies et al. 2010;Kaushal et al. 2013;Cook et al. 2020). Weed invasion is one of the major environmental degradation issues in urban waterways (Paul and Meyer 2001) and in BMUS wetlands within urbanised catchment areas (Benson and Baird 2012; Fryirs, Freidman and Kohlhagen 2012). About one-third of the BMUS wetlands are exposed to urban development, which reduces their integrity through erosion, sedimentation, modi ed hydrology and weed invasion (NSW Scienti c Committee 2007). In the current study, the contribution of concrete materials to increased growth of Salix spp. was probably enhanced by the naturally acidic and soft water of BMUS, which has underlying geology containing negligible calcium and potassium (Belmer, Wright and Tippler 2015; Carroll, Reynolds and Wright 2020; Purdy, Reynolds and Wright 2020). It is probable that urban development in landscapes containing karst geology probably has lesser geochemical impacts (water and soil) when contrasting natural versus urban catchments.
We recommend that more research is needed to investigate the contribution that concrete can have on urban water, urban soil and on invasive species in urban riparian environments. It would be bene cial to conduct further testing on how water exposed to concrete affects other invasive weed species and a range of native species, similar to the experimental exposure of phosphorus to native plants by Leishman and Thomson (2005). It would also be valuable to extend the length of plant studies for longer periods and across different plant growth stages. Further research on topics such as this will be vital in many environmental settings, as concrete is a ubiquitous urban material and will continue to be part