Archives of Environmental Contamination and Toxicology https://doi.org/10.1007/s00244-021-00827-1

The Efects of Road De‑icing Salts on Water Quality and Macroinvertebrates in Australian Alpine Areas

Mark D. Shenton1 · Susan J. Nichols1 · Jon P. Bray1 · Benjamin J. G. Moulding1 · Ben J. Keford1

Received: 1 October 2020 / Accepted: 24 February 2021 © The Author(s), under exclusive licence to Springer Science+Business Media, LLC, part of Springer Nature 2021

Abstract The application of road de-icing salts has the potential to salinize fresh waters and degrade habitat for aquatic organisms. In the Australian Alps, the ecological efects of even small salinity increases from de-icing may be diferent than in North America and Europe because of (1) diferences in the evolutionary history, and (2) areas with de-icing in Australia are not located in urbanized landscapes where de-icing has been largely studied elsewhere. In this study, we tried to determine the salinity increases attributable to de-icing in Australia and the efects of this increase in salinity to stream macroinvertebrates. We observed increased salt concentrations (as measured by continuous measurements of electrical conductivity (EC) and periodic measurements of chloride concentrations) in streams near two Australian ski resorts, during the snow seasons (June to September) of 2016 to 2018. The maximum EC observed in streams in salted sites near Perisher, New South Wales, was 390 µS cm­ −1 compared with a maximum of 26.5 µS cm­ −1 at unsalted sites. Lower EC values (i.e., maximum 61.1 µS cm­ −1) and short durations of salinity increases in streams near Falls Creek, Victoria, were not expected to cause an adverse biologi- cal response. Salt storage in the landscape was evident at salted sites near Perisher where EC was above background levels during periods of the year when no salt was applied to roads. Stream macroinvertebrate community composition difered at sites receiving run-of from road salting activities near Perisher. Abundances of Oligochaeta (worms) (up to 11-fold), Dug- esiidae (fat worms) (up to fourfold), and Aphroteniinae (chironomids) (up to 14-fold) increased, whereas Leptophlebiidae (mayfies) decreased by up to 100% compared with non-salted sites. The taxa that were less abundant where de-icing salts were present tended to be the same taxa that toxicity testing revealed to be relatively salt sensitive species. This study dem- onstrates a causal link between de-icing salts, elevated stream salinity, and altered macroinvertebrate community composition in streams that received run-of from road de-icing activity in the Australian Alps.

Salinization of freshwaters presents a growing environmen- where the total concentration of dissolved salts increase tal threat globally (Baek et al. 2014; Cañedo Argüelles et al. (Williams 1987) and is generally associated with eight major + + 2+ 2+ − 2− 2− 2013; Coldsnow et al. 2017; Fay and Shi 2012; Keford et al. inorganic ions: Na­ , ­K , ­Mg , ­Ca , ­Cl , ­SO4 , CO­ 3 , − 2016; Ziemann and Schulz 2011). Salinization is a process and HCO­ 3 (Ziemann and Schulz 2011). Salinization can have natural sources, such as groundwater discharge, geo- logical weathering, or deposition from atmospheric sources, * Mark D. Shenton [email protected] such as wind and rain (Grifth 2014; Nielsen et al. 2003). Salinization also can result from human activity, such as Susan J. Nichols [email protected] agriculture, mining, and the application of road de-icing salts in cold regions (Cañedo Argüelles et al. 2013). Jon P. Bray [email protected] The purpose of applying road de-icing salts is to lower the freezing point of water causing snow and ice to melt Benjamin J. G. Moulding [email protected] and thus reduce trafc accidents (Dailey et al. 2014). The salts produce an exothermic reaction that stimulates melt- Ben J. Keford [email protected] ing through salt dissolution and hydration. The use of road de-icing salts has increased since the turn of the twenty- 1 Centre for Applied Water Science, Institute for Applied frst century (Fay and Shi 2012). Increasing coverage of the Ecology, University of Canberra, Canberra, ACT 2601,​ landscape by roadways and increased application of road Australia

Vol.:(0123456789)1 3 Archives of Environmental Contamination and Toxicology de-icing salts are now salinizing fresh waters in cold regions Methods of North America and Europe, degrading habitat for aquatic organisms (Kaushal et al. 2005), causing acute toxicity (Fay Study Site Selection and Shi 2012), and adversely afecting community structure (Jones et al. 2017) and ecosystem processes (Tyree et al. Seven paired-site locations were selected, in subalpine 2016). There is no published empirical knowledge of the areas (hereafter alpine areas) where road de-icing salts are efects of road de-icing salts in Australia. applied to some of the roads. At each of the seven loca- The efects of road de-icing salts on salinity concentra- tions, one site was situated upstream of a road (reference tions and biota in Australian streams may difer from that site) to provide a baseline for comparison with a second in areas where it has been well studied. Mainland Australia site immediately downstream of the road (test site). Paired (i.e., excluding Tasmania) has a very small road network in sites will be referred to as site-pairs hereafter. Individual a climate where road de-icing is necessary, and these roads sites within a site-pair, either upstream or downstream of transverse mostly natural subalpine and alpine landscapes roads, will be referred to as sites hereafter. Four of the seven without signifcant urban areas. In contrast, most areas where site-pairs were located near Perisher Ski Resort within the efects of de-icing have been well studied are in adjoin- Kosciuszko National Park in New South Wales (NSW), and ing urban landscapes (Environment Canada 2001; Kaushal three were located near Falls Creek ski resort adjacent to et al. 2005). Streams within urban areas often show evidence Alpine National Park, Victoria (Fig. 1). At both Perisher of altered stream ecology, for example, reduced abundance and Falls Creek, two streams receive runof from salt de- and richness of EPT taxa, and increased abundance of dip- icing activities. Hereafter, these streams are referred to as terans and worms that are tolerant of disturbances (Walsh salted streams. The other three streams (two streams at Per- et al. 2007). Thus, the efect of salinization may be difcult isher and one at Falls Creek) received road runof from a to distinguish from those of urbanisation. road with no recent application of salt, because the road is The Australian Alps are within largely natural and non- closed in winter. Hereafter, these streams will be referred urbanized systems, and the streams are characterised by to as nonsalted streams. The purpose of sampling at site- extremely low salinity water (Williams et al. 1970). Thus, pairs on the nonsalted streams was to provide a control or the organisms living in this environment may be adapted baseline with which to compare samples from salted site- to living in very dilute water and are potentially unable to pairs, thus allowing greater confdence when linking results tolerate elevated salinities (Keford 2019) and/or are vulner- to road de-icing salts rather than any efect of roads (e.g., able to being outcompeted if salt tolerant species are present from pollutants coming from vehicles). At Perisher Creek, (Bray et al. 2019). Furthermore, previous studies indicated a third site was located approximately 2 km downstream of the aquatic macroinvertebrate assemblages in the Austral- winter de-icing activities to test the extent of efects from ian Alps have a high proportion of salt sensitive taxa, such any salinity increases. See supplementary data Figs. S1 and as Ephemeroptera (mayfies), Plecoptera (stonefies), and S2 and Tables S1.1 to S2.2 for locations of individual sites Trichoptera (caddis fies) (EPT), and some of the groups and site-pairs. from these orders have a diferent evolutionary history to those found in Northern Hemisphere regions (Clements et al. Macroinvertebrate Field Sampling and Laboratory 2016). In summary, non-Australian studies of efects of de- Processing icing salt efects on stream macroinvertebrates may plausibly be of limited use for determining the efects in Australia. Benthic macroinvertebrate sampling was undertaken in The paucity of knowledge regarding salinity increases spring/summer (November to December) and autumn (May) attributable to road de-icing activities in Australia, and in 2016 and 2017. Spring samples were collected from each environmentally acceptable salinity levels in Australian site when macroinvertebrates had minimal time to recover alpine streams, makes it difcult for managers to practice following any efect of the de-icing salts over the winter evidence-based management of snow and ice in Australia. snow season. Samples collected in late autumn were taken The purpose of this paper was to determine the efect of when macroinvertebrates are expected to have had maxi- road de-icing on salinity in Australian alpine streams and mum time to recover from any winter salinity peaks asso- the macroinvertebrate communities. ciated with road de-icing. Two macroinvertebrate samples were collected from each site during each sampling episode, with one selected at random for identifcation because of project time constraints. Macroinvertebrates were sampled from rife habitats following AUSRIVAS sampling proto- cols (Nichols et al. 2000). In brief, macroinvertebrates were

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Fig. 1 Study locations relative to Canberra, ACT (35° 16′ 55.2036′’ S, 149° 7′ 44.3928′’ E). Study sites are located near Perisher, NSW; and near Falls Creek, Victoria

collected from a 10-m section of rife habitat using a 350- predict whether site-pairs were expected to have the same mm D-shaped kick net with 250 µm mesh, while disturbing or similar macroinvertebrate assemblages in the absence of the stream substrate to a depth of approximately 100 mm. any efect of the roads or de-icing. Preserved samples were then subsampled (Marchant 1989) The habitat factors that the AUSRIVAS model uses were in the laboratory until approximately 200 individuals were chosen by its developers after sampling many reference sites identifed (Nichols et al. 2000). Macroinvertebrates were and measuring many habitat factors at each site (Simpson identified to family taxonomic level with the exception and Norris 2000). In developing the model, habitat variables of Oligochaeta (worms) and Acarina (mites), which were that best distinguish between groups of reference sites (i.e., identifed to class, and Chironomidae that were identifed reference sites grouped on the similarity of their macroin- to subfamily using online identifcation keys available from vertebrate assemblage) were determined. These variables the Murray-Darling Freshwater Research Centre (MDFRC) are then the variables that the model used to predict the (https://​www.​mdfrc.​org.​au/​buggu​ide/​displ​ay.​asp?​type=​1&​ macroinvertebrate community at sites with similar habitat class=​19). This taxonomic resolution was chosen for con- (and are the variables we used to run the model). What we sistency with AUSRIVAS predictive model requirements did in the current study was to compare the communities (see below). predicted by the AUSRIVAS model between the site-pairs as a check that the macroinvertebrate community was likely to AUSRIVAS Predictive Models have a similar macroinvertebrate community in the absence of the road and/or salting. Specifcally, we used an AUSRI- AUSRIVAS is predictive modelling software that predicts VAS model developed for that region (Nichols et al. 2010) to the taxa expected at sites in the absence of environmental compare the number and type of taxa predicted to occur with disturbance (Simpson and Norris 2000). The AUSRIVAS a ≥ 50% probability at upstream and downstream sites within model was used because the paired survey design has an site-pairs to ensure the assumption of similarity between implicit assumption that the macroinvertebrate community sites is satisfed. would be the same or similar at sites upstream and down- The habitat variables used in the predictive model were stream of the roads at the selected locations. Macroinverte- substrate variables (percentage of cobble, boulders, and brates are afected by many abiotic factors (e.g., Naden et al. detritus), habitat scores, velocity-depth scores, water tem- 2016; Richards et al. 1993). Thus, we used AUSRIVAS to perature, location variables (latitude, longitude, and distance

1 3 Archives of Environmental Contamination and Toxicology from the source), percentage of macrophytes present, ripar- or near, study sites upstream and downstream of salted roads ian variables (canopy cover at the river and vegetation between June to October at least monthly in 2016, 2017, cover), bottom scouring and deposition scores, and pool- and 2018. rife-run-bend ratio (see Nichols et al. (2000) for details on how these are measured). Anion Composition

− 2− − − − − Continuous Salinity and Stream Depth Major anion content ­(Cl , ­SO4 , ­F , ­Br , NO­ 2 , and ­NO3 ) Measurements of the water samples was determined using ion chromatog- raphy (IC). Anion analysis methods follow those detailed in As a result of the short and pulsed nature of runof expected United States Environmental Protection Agency (US EPA) with snow-melt events (Coldsnow et al. 2017), periodic Method 300.0 (Jackson 2001) and Dionex Application Note measurements of stream salinity are highly likely to severely 133 (Jackson 2015). Samples of NaCl and CaCl­ 2 road de- underestimate peak salinity concentrations. Thus, electri- icing salts were obtained from Transport NSW contractors cal conductivity (EC) data loggers (HOBO U24-001) were and analysed for major anions as described above for the deployed at each site on the salted-road streams between purpose of identifying content of anions contributed to May 2016 and May 2019 to collect EC (µS ­cm−1 @ 25 °C) waterways from road de-icing operations. In addition to the and water temperature (°C) data at 15-min intervals with methods outlined above for IC analysis of samples, calibra- 83% of the timeseries continuous. Loggers were not tion standards and ultrapure water blanks were used as qual- deployed at some paired sites on nonsalted streams, because ity control and quality assurance measures when analysing they were located near roads that are closed for the dura- the de-icing salt and water samples. The levels of ­Cl− meas- tion of the snow season. Therefore, no EC diferences were ured are not an indication of the total efect of salinity, expected relative to paired-sites on salted streams located because some of the efect will come from the cations (e.g., upstream of salted roads where loggers had been deployed. Mount et al. 2016). ­Cl− levels are reported to allow com- Only minor diferences between paired sites on nonsalted parison with studies reporting salinity in terms of that anion. streams might be expected to occur through relative difer- ences in evapotranspiration. Before any outlier datapoints Precipitation and Road De‑icing Salt Application were removed, they were checked against Bureau of Mete- orology (BoM) (see www.bom.​ gov.​ au​ ) precipitation records, Precipitation data was obtained from the Bureau of Mete- depth logger data, and water temperature data to determine orology (BoM) (www.​bom.​gov.​au) stations at Perisher and whether a satisfactory explanation could be found. Data- Falls Creek. Road de-icing salt application records were points were only deleted if cross-checks could reasonably obtained from contractors to NSW RMS for the 2017 snow demonstrate that an error had occurred; otherwise, they are season and from the Falls Creek Resort for the 2015 to 2017 assumed to be real datapoints and remain in the dataset. snow seasons. These data were analysed to determine rela- Note: An EC logger was lost from the Perisher Carpark site tionships between precipitation and transport of road de- in 2018 containing winter data, and technical issues pre- icing salts in the environment. vented collection of EC data at the Pipers Downstream site (supplementary data, Fig. S1) during the 2018 ski season. Rapid Toxicity Tests The EC values obtained by logger readings at Falls Creek site-pairs were not at levels or durations expected to cause an Rapid toxicity tests (RTT) were used to determine the salin- adverse change in the macroinvertebrate community; there- ity sensitivity, under laboratory conditions, of many mac- fore, investigation of macroinvertebrate response to elevated roinvertebrate species taken from specifc communities of salinity did not proceed at Falls Creek. interest so that we could compare the species sensitivity distributions (SSDs) (Kon Kam King et al. 2015) of (a) taxa Periodic Water Sample Collection with diferent apparent responses in the feld study and (b) taxa collected from an alpine area to those from low-land Two fltered water samples (0.45 µm, Sartorius Stedim Bio- eastern Australian rivers. tech) were collected concurrent with the aforementioned macroinvertebrate samples. EC, water temperature, pH, Field Collections for RTT​ and dissolved oxygen also were measured in situ using a Horiba U-50 series multiprobe, and alkalinity was meas- Alpine stream macroinvertebrates were collected between ured by titration. Additionally, two fltered (0.45 µm) water May and September 2018 from six sites at 1,400 to samples, EC, water temperature, pH, and dissolved oxygen 1,600 m ASL near Perisher, NSW, all within Kosciuszko readings were semiregularly collected from sites located at, National Park (refer Supplementary data, Table S2). Water

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temperature during collections ranged from 0.4 to 5.2 °C. To determine a ­LC50 for salinity for each taxonomic Invertebrates were kept under aeration in site water with group that were found in low abundance (e.g., 1–2 individ- some detritus and sediment during transport to the labo- uals per test), up/down tests were performed on the same ratory. Salinity toxicity tests commenced within 24 h of taxa (Keford 2014). For example, if taxa from a particular collection. group survived at “X concentration,” then in the next test the same taxa were placed into a treatment “ > X concen- Laboratory Processing and Toxicity Testing tration.” If mortality was seen in the higher concentration compared with the lower concentration, a lethal concentra- In the laboratory, separate taxonomic groups were ini- tion range for that taxon was given. Therefore, dependent tially identifed by eye to avoid damage/stress under a on the RTT data, lethal concentration values are given as microscope and were placed into tea balls (stainless steel a point estimate, a range, or greater than. strainers designed for tea making) within aquariums that were dosed with 3 L of predetermined concentrations of artifcial marine salts (AMS) (Ocean Nature brand), with Data Analyses concentrations determined from previous lethal concentra- tion (LC) values in the literature (e.g., Keford et al. 2003, Macroinvertebrate Community Composition 2012). A maximum of 10 animals of the same taxon were placed in each tea ball, with particular attention paid to Bray–Curtis similarity coefcients (Bray and Curtis 1957) separating predatory taxa from other taxa. AMS was used were calculated for macroinvertebrate abundance data to allow comparison with pre-existing salt sensitivity data using PRIMER6 (Clarke and Gorley 2006) for relevant for Australian lowland stream macroinvertebrates, all of site-pairs and values manually extracted from the similar- which was obtained using AMS as the salt source (e.g., ity matrix. Sites with identical macroinvertebrate assem- Keford et al. 2003, 2012), so we could determine whether blages have a Bray–Curtis similarity coefcient of 1, with this pre-existing data could be validly used to inform the coefcient values moving toward 0 as assemblages become likely efects of increased salinity in alpine areas of Aus- less similar in composition. Coefcient values were then tralia. This was important, because we hypothesised that tested to determine whether macroinvertebrate assem- the Australian alpine fauna may be more salt sensitive than blages from salted sites difered from unsalted sites using lowland fauna, meaning salinity guidelines developed for a linear mixed-efects model ANOVA (R-Studio version lowland regions of Australia might not be validly appli- 1.1.453, R version 3.5.0) (α = 0.05) where the paired sites cable to alpine or subalpine areas of Australia. Controls were set as the random factor and salted (salted stream and water used for dilution solutions was dechlorinated sites or unsalted sites), season (autumn or spring), and Canberra tap water. There was no acclimation period for year (2016 or 2017) were fxed efects factors each with the taxa before being placed into treatments. Little to no two levels. Assumptions of normality were examined using replication of treatments are used when conducting RTTs, a histogram of residuals and quantile–quantile (Q–Q) plots because in most collections, there are not enough indi- (Wilk and Gnanadesikan 1968). Assumptions of equal var- viduals to allow replication (Hickey et al. 2008; Keford iance were examined using residual versus ftted values et al. 2005a). All results reported here are for tests last- plots. ing 72 h, and all treatments were aerated with air stones Nonmetric multidimensional scaling (nMDS) was used with no further substrate added. No individuals were fed to visualise diferences in macroinvertebrate assemblages during the tests, which were performed in temperature- at each site using Hellinger transformed (Legendre and controlled cabinets set at 7.0 °C ± 1.0 °C, with 8 and Gallagher 2001) macroinvertebrate abundance data, and 16 h day and night, respectively. All toxicity tests used ordinations were plotted using Bray–Curtis distances. Vec- the standing water system (Keford et al. 2005b), where no tors were generated by the “vegan” package in R-Studio. further infow of water was added and fow is provided by Site physical and chemical data correlation relationships aeration. All aquaria were inspected daily, recording the were then explored and overlaid on the ordination using survival of individuals, and any expired individuals were the R function “envft.” All backwards stepwise selected removed from treatment aquariums as soon as found. Taxa variables were correlated (i.e., > 0.7). identifcation was confrmed post-mortem by microscopy Permutational ANOSIM (ANalysis Of SIMilarity) (999 using the same identifcation keys discussed earlier. Mor- permutations) using the “vegan” package in R-Studio was tality was defned as not moving and failing to respond to undertaken to examine data for overall signifcant difer- probing with a blunt instrument or gentle squirt of water ences between macroinvertebrate communities at salted from a pipette. Dead individuals and those surviving at the and unsalted sites using relative abundance data (i.e., end of the test were preserved in 80% ethanol. % composition). This was followed up with SIMPER

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(SIMilarity PERcentage) analysis to identify taxonomic Australian streams obtained from published literature groups that made a significant contribution to those (Kefford et al. 2003, 2012), and SSDs for Australian diferences. alpine taxa obtained in this study, to determine whether LC­ 50 values from lowland eastern Australian streams can be used as a proxy for LC­ 50 values for Australian Chloride Concentration Data alpine taxa. Further comparisons were then made within the Australian alpine taxa group by generating SSDs Water samples collected from, or nearby to study sites, for those taxa which, irrespective of statistical signifi- were analysed for chloride concentration content. Values cance, appeared to decrease, increase, or had no apparent obtained were statistically analysed by ANOVA using a response to salinity inputs. mixed effects model with factor site as the random fac- tor, and factor salted as a fixed effect factor with two levels (salted and unsalted). Chloride data were log(x + 1) Results transformed to meet the assumption of equal variance. Assumptions of normality were examined using a his- Efect of Road De‑icing on Stream Salinity togram of residuals and quantile–quantile (Q-Q) plots. Assumptions of equal variance were examined using Diferences were detected in EC between sites upstream residual versus fitted values plots. and downstream on Australian alpine roads that had de- icing salts applied, with diferences most notable at site- pairs located in the Perisher, NSW region. At Perisher sites Species Sensitivity Distributions upstream of salted roads, salinity concentrations (measured as EC) were always low (mean 8.8 µS ­cm−1 ± 0.01 SE, range Species sensitivity distributions were generated by the 1.1–26.4 µS cm­ −1). However, at Perisher sites downstream MOSAIC­ SSD web-interface (Charles et al. 2018), using of salted roads, mean salinity concentrations were approxi- 72-h LC­ 50 values generated via RTTs. The 72-h ­LC50 mately double those relative to upstream sites (mean 17.3 values were used to compare the salt sensitivity of the µS ­cm−1 ± 0.03 SE, range 0.3–389 µS ­cm−1). Salinity at all Australian alpine and lowland macroinvertebrates because Falls Creek sites remained relatively low throughout the year available data for the lowland macroinvertebrates is pre- (mean 10.0 µS cm­ −1 ± 0.01 SE, range 0.3–61.1 µS cm­ −1, dominantly over 72 h of exposure. These were then plot- inclusive of downstream sites receiving salt runof). ted using a log-normal distribution. Comparison of SSDs A salinity “peak” hereafter is defned as any increase in were then made between taxa found in lowland eastern salinity ≥ 40 µS ­cm−1. This value was chosen so that peaks

Table 1 Salinity peaks (EC ≥ 40 µS ­cm−1) observed at sites receiving salt runof from roads Site Year # Peaks Max peak EC Mean peak EC Max dura- Mean duration (h) ± SE (µS ­cm−1) (µS ­cm−1) ± SE tion (h)

Perisher carpark 2016 8 101 56.6 ± 1.5 4.8 2.7 ± 0.5 Perisher carpark 2017 35 247 71. 0 ± 0.7 146 15.7 ± 5.1 Perisher carpark 2018* 1 59.4 50 ± 1.7 2.5 NA Perisher carpark 2019 1 45.6 43.4 ± 1 0.8 NA Perisher carpark All years 45 247 62.1 ± 0.7 146 12.8 ± 4.1 Perisher Creek 2 km downstream of carpark 2016 4 74.2 50.8 ± 0.2 2.3 1.4 ± 0.5 Perisher Creek 2 km downstream of carpark 2017 27 131 60.7 ± 0.5 114 11.1 ± 4.2 Perisher Creek 2 km downstream of carpark 2018 23 126 52.1 ± 0.3 136 20.0 ± 6.6 Perisher Creek 2 km downstream of carpark 2019 0 NA NA NA NA Perisher Creek 2 km downstream of carpark All years 54 131 55.5 ± 0.3 136 14.5 ± 3.6 Pipers downstream 2016 16 145 55.4 ± 0.8 72 8.7 ± 4.4 Pipers downstream 2017 42 389 75.7 ± 0.5 511 26.7 ± 12.5 Pipers downstream 2018* 29 251 70.5 ± 0.7 151 51.4 ± 4.0 Pipers downstream All years 87 389 72.2 ± 0.4 511 24.4 ± 6.8

“Year” refers to the entire calendar year. Entries denoted * do not contain data for the 2018 snow season. 2019 data only includes measurements before the 2019 snow season (i.e., these peaks occurred before commencement of the snow season in June 2019)

1 3 Archives of Environmental Contamination and Toxicology are those EC values that exceed approximately double the Chloride concentrations of stream water collected highest EC value observed at upstream sites over the period from unsalted sites did not exceed 5 mg ­L−1 with a mean of the study (26.4 µS ­cm−1) and for a duration of 30 min of 1.75 mg L­ −1 ± 0.1 SE. Mean chloride concentrations at or more (Table 1). Otherwise, the term “pulse” is used to salted sites were 15.42 mg ­L−1 ± 6.4 SE (refer supplemen- describe any increase in EC above background levels, irre- tary data, Tables S5.2 and S5.3). Diferences in mean chlo- spective of duration. ride concentrations of water samples between salted and Salinity peaks mostly occurred during the snow sea- unsalted study sites located near Perisher are statistically son (June to September) (refer supplementary data Fig- signifcant (t = 7.02; df = 7; p = 0.004). No statistically sig- ures S3.1–S3.4), although some salinity peaks also occurred nifcant diferences were demonstrated for sulfate or fuoride at Perisher sites outside of that period (i.e., October to the concentrations (p = 0.08 and 0.33 respectively). All water following May). Peaks outside of the snow season gener- samples from sites in the Perisher study area containing ele- ally reached a lower maximum (maximum 250 µS ­cm−1 at vated chloride concentrations were taken from sites receiv- Pipers Downstream in 2017/18) (refer supplementary data ing runof from salted roads. The highest chloride concen- Table S5), but these peaks were smaller. In most cases, salin- tration (360 mg ­L−1) was found in a water sample collected ity peaks occur within 24 h of precipitation (refer supple- during the 2017 snow season from the Pipers Downstream mentary data Figures S2.1–S2.4). Dilution of stream water site. Plots of chloride concentrations can be found in sup- by a mean value of 24% (averaged across the entire time plementary data, Figures S3.1 and S3.2. series) was evident from the downstream site Perisher car- park to the far downstream site Perisher Creek 2 km down- Macroinvertebrate Assemblage Structure stream of carpark (refer supplementary data, Figure S3.4; Table S5), which are located approximately 2 km apart . A total of 53 macroinvertebrate taxa were collected in this study. 87% of the individuals were insects. Across all sites Major Anionic Composition of De‑icing Salts the most abundant taxa were Gripopterygidae, Oligochaeta, and Stream Water Elmidae, Conoesucidae, and Orthocladiinae. EPT taxa com- bined comprised a mean 47% of individuals in the samples Chloride content of NaCl and CaCl­ 2 salt samples used in (SD ± 18%, range 9.8–75%), averaged across the entire NSW road salting operations as a percentage (%) of total dataset. anions (by mass) were 99.88% and 99.15%, respectively. AUSRIVAS modelling showed that taxa predicted to Bromide, nitrate, and sulfate made up the remainder of the occur with a ≥ 0.5 probability were the same or similar (± 2 detectable major anion content of the road salt samples taxa maximum) between upstream and downstream site- (< 0.1% total). Fluoride, nitrite, and phosphate content were pairs (supplementary data, Table S6). We are therefore con- below detectable limits (i.e., < 0.0001%, refer supplementary fdent that any diferences observed in macroinvertebrate data, Table S5.1). assemblages between the site-pairs is the result of either the efect of roads or salt applications.

Fig. 2 Bray–Curtis similarity coefcients from salted site- pairs to unsalted site-pairs in 2016 and 2017. Salted site-pairs are less similar in taxonomic composition than unsalted site-pairs in both years. Each site-pair has an individual site both upstream and downstream of a road. Note: Lines in centre section of boxplots indicate the arithmetic mean. Boxed sections show the interquartile range

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Bray–Curtis coefcients by factors Year and Salted show Table 2 Correlated physicochemical variables from nMDS analysis site-pairs at salted streams are less similar in assemblage of relative macroinvertebrate abundance between salted and unsalted sites (999 permutations) structure than site-pairs at unsalted streams, and that this was the case in both 2016 and 2017 (Fig. 2). The greatest dif- Vector Defnition R2 p value ference in mean Bray–Curtis similiarity between site-pairs SpotEC EC measurement at time of sampling (mS 0.33 0.003 at salted and unsalted streams occurred in 2016. SIMPER ­cm−1) analysis of relative abundance data (i.e., % composition) AvgEC Mean EC over sampling season (µS ­cm−1) 0.24 0.007 between site-pairs at salted and unsalted streams found sig- MaxEC Maximum EC over sampling season (µS 0.17 0.042 nifcant diferences between Oligochaeta (p = 0.001), Lep- ­cm−1) tophlebiidae (p = 0.027), Aphroteniinae (p = 0.045), and MaxCl Maximum ­Cl− concentration (mg ­L−1) 0.14 0.049 Dugesiidae (p = 0.005). Statistical signifcance (p > 0.05) was not observed for any other taxa. Oligochaeta, Aphrote- niinae, and Dugesiidae abundances were greater at salted of the plot rather than randomly spread across the ordination sites than unsalted sites. Leptophlebiidae abundances were space (Fig. 3; Table 2). nMDS weighted average site scores less at salted sites than unsalted sites (refer supplementary are provided in supplementary data, Table S8. data, Figures S4.1–S4.4 inclusive). Considering all sites separately, i.e., not as site-pairs, per- Species Sensitivity Distributions mutational ANOSIM demonstrated a small, but statistically signifcant diference in the macroinvertebrate assemblage Australian alpine taxa were more sensitive to salinity (in structure at sites receiving salt runof (salted sites) and those terms of 72-h ­LC50 values) than those from Australian low- sites not receiving salt runof (unsalted sites) (R = 0.167; land streams (Fig. 4). 72-h ­LC50 values obtained in this study p = 0.007). The ANOSIM output plot is provided in sup- are provided in supplementary data, Table S7. plementary data, Fig. S5. Taxa that appeared to decrease, showed no efect, or increased at salted sites (regardless of statistical signif- Nonmetric Multidimensional Scaling (nMDS) cance) had diferent 72-h ­LC50 values (Fig. 5). The taxo- nomic group that decreased in abundance (decreasing taxa) Two-dimensional nMDS plot (Fig. 3) (stress = 0.25) of sta- were Baetidae (Ephemeroptera), Chironomidae (Diptera), tistically signifcant factors determining the composition and Leptophlebiidae (Ephemeroptera), and they were more of macroinvertebrate communities show a pattern along salt sensitive ­(LC50 values range from 4 to > 15 mS/cm) than nonmetric multidimensional scaling (nMDS) axis 2, where taxa from the “no efect” ­(LC50 range > 15 to 35 mS/cm) and salted sites with greater EC are located within the upper part “increasing taxa” ­(LC50 range < 5 to > 30 mS/cm) groups.

Fig. 3 Two-dimensional nMDS of macroinvertebrate commu- nity data (stress = 0.25). Blue dots represent salted sites; black dots represent unsalted sites. The size of dots is relative to maximum EC recorded at each site for the season in which it was collected: large dots are higher EC; small dots are lower EC. Correlated physicochemical variables overlaid on this ordi- nation are defned in Table 2

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Fig. 4 Alpine taxa SSD (dashed line) compared to lowland taxa SSD (solid line) based on 72-h ­LC50 values. Gray boxes repre- sent regions of non-uniqueness in the non-parametric maximum likelihood estimates of the empirical cumulative distribu- tion functions (NPMLE ECDF). See Wang and Taylor (2013) for details

Fig. 5 Comparison of SSDs for 72-h LC­ 50 values of taxa that appeared to decrease in response to salinity (a Decreasing Taxa), to those that appeared to have no response to salinity increase (b No Efect Taxa) and those that appeared to increase in response to salinity increases (c Increasing Taxa). Gray boxes represent regions of nonuniqueness in the NPMLE ECDF (see Wang and Taylor (2013) for details)

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Discussion In light of the high conservation value and the normally dilute nature of Australian alpine streams (Williams et al. Salinity Increases from Road Salting 1970) relative to those found in other regions (e.g., Kaushal et al. 2005), and the changes seen in macroinvertebrate com- During the snow season substantial increases in EC and munities observed in this study, we argue that the current −1 Cl­ − concentrations occurred at sites downstream of salted guideline trigger value of 350 µS cm­ in ANZECC and roads relative to their upstream pairs at Perisher, NSW. ARMCANZ (2000) for NSW upland waters is not biologi- Increases in EC and Cl­ − at levels expected to cause an cally meaningful or protective. Based on the highest EC −1 adverse biological response were not observed at Falls measurement observed in the current study of 389 µS ­cm , Creek, Victoria. Sites not receiving runof with de-icing a lower EC trigger value in line with other Australian states −1 salts always had very low EC values and Cl­ − concentrations (e.g., 30–120 µS cm­ in Victoria for upland waters), or (refer supplementary data, Tables S4 and S5; Figs. S3.1 and development of an EC trigger value specifc to NSW alpine S3.2). At sites downstream of salted roads at Perisher the waters, would be more suitable in determining a manage- peak EC values were 4- to 10- (in 2016), and 10- to 15-fold ment response for the protection of Australian alpine stream (in 2017), greater than at upstream unsalted sites (refer biota. Table 1). Downstream salted sites saw ­Cl− concentrations up Although much lower and less frequent than during the to 360 mg ­L−1, with a mean of 15.4 mg L­ −1. Peak EC levels salted season, elevated EC was detected outside of the snow (250–390 µS ­cm−1) at downstream salted sites are at levels seasons (October to the following May) at NSW sites down- with the potential to cause adverse biological responses in stream of salted roads. Mean EC values at salted sites during stream macroinvertebrate communities (Clements and Kota- the unsalted season were up to double those observed at lik 2016; Keford et al. 2011; Olson and Hawkins 2017). unsalted sites during 2016/17 and 2018/19, and up to triple For example, EPT richness was 25% less at 200–300 µS upstream EC values during 2017/18. The increase in salin- cm­ −1 compared with < 50 µS ­cm−1 (Keford et al. 2011), ity outside of the snow season indicates that salts are being and abundance of baetid mayfy and total mayfies, com- stored in the landscape from winter salting activities, or that munity metabolism and taxa richness were reduced while there is a lag efect lasting several months or more, which is mayfy drift increased at, or near to 300 µS cm­ −1 (Clements raising salinity levels over the entire unsalted season (Octo- and Kotalik 2016; Olson and Hawkins 2017). The current ber to the following May). The current study is unable to study found that salted site-pairs are less similar in commu- provide an estimated timeframe for salt storage in the land- nity structure than unsalted site-pairs, and that these difer- scape, but estimates for other regions range from months ences are attributable to reduced Leptophlebiidae (mayfy) to years (Environment Canada 2001). The amount of salt abundance, and increased Oligochaeta (worms), Dugesiidae running directly into Australian alpine waterways relative (fat worms) and Aphroteniinae (chironomids) abundances. to that remaining in the landscape is unknown and cannot Moreover, (disregarding statistical signifcance) those taxa be currently estimated. Our study shows that salt inputs are that tended to decrease at salted sites, toxicity tests indi- still detected before the snow season begins, approximately cated that they were more salt sensitive than taxa that had 8 months after de-icing operations from the previous snow no change in their abundance or increased in abundance at season have ceased. This indicates that salt storage is occur- salted sites (Fig. 5). For Australian alpine regions, where ring for at least 8 months. EPT taxa normally comprise the majority of taxa within Elsewhere, the efects of de-icing salts have typically been stream macroinvertebrate communities, there is the potential studied in urbanized environments where much of the water for de-icing runof to afect a large proportion of taxa found catchments are covered by surfaces impervious to water in these regions (Suter et al. 2002; Campbell et al. 1986, (e.g., roads, carparks, and pavements). These studies indi- both cited in Clements et al. 2016). cate that where de-icing salts are applied, they are strongly Australia has not set chronic chloride concentration val- associated with salinity levels and biological impairment in ues for the protection of stream biota, but the maximum streams (Corsi et al. 2010; Wang and Kanehl 2003; Wang chloride concentrations observed in this study (360 mg et al. 2001). For example, the U.S. chronic water quality cri- − ­L−1) exceed those set in the United States (230 mg ­L−1) and terion for ­Cl is exceeded beginning at approximately 10% Canada (120 mg L­ −1). Considering that water samples for impervious surface coverage (Corsi et al. 2015; Cufney Cl­ − was obtained approximately monthly during the snow et al. 2010). The current study, unlike most fndings in the season and that EC values varied by as much as 350 µS ­cm−1 United States, observed elevated salinity with < 1% impervi- within a day, the maximum observed ­Cl− concentration is ous surface coverage of the catchment. highly likely to be an underestimate of the actual maximum The severity of salinization caused by leaching and Cl­ − concentration in the streams downstream of salted roads. groundwater inputs will vary according to stream fow but is expected to be more severe in summer months and drought

1 3 Archives of Environmental Contamination and Toxicology years, particularly following snow seasons where substantial is important, because diferent ionic proportions are known amounts of salts were used. Less fow in streams outside of to have diferent efects on stream macroinvertebrates, even the snow season can reduce dilution capacity, and in many when EC remains similar (Cañedo-Argüelles et al. 2016; instances occurs during times of year when earlier and more Kunz et al. 2013; Zalizniak et al. 2006). For example, sub- sensitive life-cycle stages of aquatic macroinvertebrates are lethal testing by Zalizniak et al. (2006) showed the greatest in greater abundances. For example, Keford et al. (2004) macroinvertebrate response in water types containing little found that eggs and hatchlings were often more salt sensitive or no ­Ca2+ cations. Surface waters with naturally low ionic than older life-stages, whereas Camp et al. (2014) indicated proportions, such as those found in Australian alpine waters, susceptibility to toxicants during the moulting of insects. are those where we expect taxa to be most sensitive to small This has particular relevance to mayfies, which can have up salt inputs. Additionally, macroinvertebrates are likely to be to 52 larval moults (Gros 1925, cited in Camp et al. 2014) more sensitive to NaCl than to AMS (Keford et al. 2005b). compared with a typical four or fve moults in other groups Therefore, the toxicity results obtained in this study (Figs. 4 of freshwater insects (Benke and Wallace 1980). Thus, and 5) are likely to underestimate true toxicity values. elevated salinity outside of the snow season is expected to The Pipers Creek downstream site consistently had the afect the ability of salt-sensitive taxa to recover from epi- highest EC and ­Cl− readings, as well as the greatest biologi- sodic salinity related disturbances, particularly those taxa cal impairment of any site sampled. This can be explained that have very low salinity tolerances, long generation times by a combination of factors. First, the downstream study (≤ 2 generations per year), and limited dispersal abilities site is located in an area that runs parallel to Kosciuszko (e.g., some mayfies). Road for approximately 1 km, so it is likely to be receiv- Increases in EC were typically seen as pulse-type events ing more de-icing salt runof relative to other study sites throughout the year, based on a time scale relative to the because of the exposure to larger surface area receiving snow season. The highest EC always occurred within 24 h of road de-icing salts. There is a strong relationship between precipitation. Mean peak duration (observed across all sea- ­Cl− concentration in streams and length of the roadway that sons and years) was 9 h; however, the longest peak continued drains directly into the watercourse (e.g., where the road is for 21 days. Extended peaks in EC occurred, because EC had adjacent to the watercourse) (Environment Canada 2001; not returned to background levels before a new pulse arrived. Mattson and Godfrey 1994). Second, the longer road length This was particularly evident during the snow season. In being salted provides more area for salt to be stored in soils addition to pulse disturbances, there was evidence of press- and groundwater because the absence of gutters allows salts type disturbances, in which EC readings were raised above to run directly into the adjacent roadside environment. Third, background levels, typically for 4–8 days at downstream additional salts that are applied in the carpark area also drain salted sites during the snow season and 7–24 days outside into Pipers Creek between the upstream unsalted site and the of snow seasons. Interactions between pulse and press-type downstream salted site. Studies in other regions have shown disturbance have been shown to afect the magnitude and a strong link between urbanized and built-up environments duration of macroinvertebrate communities’ responses (Col- and salt runof because of increased impervious surface cov- lier and Quinn 2003). erage (Daley et al. 2009). Fourth, Pipers Creek is a smaller At the one stream (Perisher Creek) where downstream waterway compared to Perisher Creek and therefore has less dilution was measured 2 km downstream of road de-icing capacity to dilute salt inputs throughout the year. activities, EC had decreased by a mean 24% (averaged across There are three possible explanations for the lack of the entire dataset) from its value at a site directly below the increase in salinity at Falls Creek, which may be operating road and carpark where maximum salinity levels would have individually or in combination. First is the relatively small occurred. How much further downstream for Perisher Creek, amounts of road de-icing salts used by Falls Creek Resort, and other salted streams, to return to ambient salinity lev- up to 20 tonnes annually (personal communications with els is unknown. Each catchment, however, will be diferent Jessica Millet-Riley and Geofrey Sorensen of Falls Creek because of variation in site hydrology and stream dilution Resort) compared with an estimated 890 tonnes near Per- capacity and requires assessment on a case-by-case basis. isher during the 2017 snow season (personal communication The conclusion drawn from the current study is that elevated with Adrian Walsh of Transport NSW). Second, most of the salinity from road de-icing salts may extend at least two 20 tonnes of road salts applied at Falls Creek are to carparks kilometres downstream. and roads within Falls Creek Resort area, which are located This study found no statistically signifcant change in the outside of the catchments of the streams we studied. A third concentrations of fuoride and sulfate anions from unsalted factor is the very steep topography of the study sites in the to salted sites but did fnd a signifcant diference in the pro- area surrounding Falls Creek, which may be causing some portions and concentrations of chloride anions (a 13-fold de-icing salts to run of roads directly down the hillslope increase). The specifc ionic composition of stream water

1 3 Archives of Environmental Contamination and Toxicology without entering the stream until further downstream of our salinity in Australian alpine areas. The taxonomic grouping salinity loggers and therefore not detected by our study. of EPT appears to be too broad, rendering it insensitive to changes, because although we observed declines or absence of Ephemeroptera during 2017 when EC pulses were higher Efects on Macroinvertebrate Community and for longer periods of time, the Plecoptera and Trichop- tera family richness increased at some sites. Road salting at Perisher is adversely affecting stream Although no statistically signifcant diferences were macroinvertebrate communities at salted sites. Site-pairs found in the number of individuals within order Ephemer- upstream and downstream of salted roads are less similar in optera between salted and unsalted sites overall, these taxonomic composition compared to site-pairs at unsalted taxa were absent from samples taken at the Pipers Down- roads. The taxa mostly contributing to changes in communi- stream (salted, downstream site) throughout 2017, despite ties were increases in the densities of Oligochaeta (worms), comprising 8–12% of individuals at the upstream site-pair Dugesiidae (fat worms), and Aphroteniinae (chironomids) (Pipers Upstream) for the same period. In 2016, these at salted sites and decreases in Leptophlebiidae (mayfies) at taxa accounted for < 2% of individuals found at the down- salted sites. These changes are consistent with changes seen stream site but were present at 11–20% of the individuals elsewhere (Environment Canada 2001; Kersey and Mac- at the upstream site. Ephemeroptera were not recolonising kay 1981; Williams et al. 1999). However, the EC levels at the downstream site (e.g., through downstream drift of observed in this study are lower than those often reported approximately 500 m), suggesting unfavourable conditions in other regions, which may indicate lower EC thresholds for Ephemeroptera. In contrast, downstream recovery of for changes to macroinvertebrate communities in Austral- this group is evident from Perisher carpark (downstream) ian alpine regions, relative to elsewhere. Such diferences to Perisher Creek 2 km downstream of carpark (further in tolerance could plausibly occur because of the extremely downstream). Additionally, Leptophlebiidae abundance dilute waters, high proportion of EPT taxa, and more natural was statistically signifcantly lower at salted sites relative to environment found in Australian alpine areas. upstream sites. The decrease in Leptophlebiidae abundance Salinity levels were higher and remained elevated for appears not to be attributed to changes in their food, peri- longer at all downstream sites receiving salt laden runof in phyton (Gooderham and Tsyrlin 2002), as decreases in other 2017 compared with 2016. Thus, we would have expected families from the same functional feeding group (scrapers), less EPT taxa in 2017 than 2016; however, overall the such as Elmidae, Conoesucidae, and Gripopterygidae (Goo- reverse was observed. The EC increases observed at salted derham and Tsyrlin 2002) were not observed. Other factors, sites in this study were within the range in which increases including substrate, pH, stream width and depth, and other in family richness are expected while Ephemeroptera simul- habitat variables, all appeared to be favourable. Although taneously decline (Keford et al. 2011). EPT taxa are gener- speculative, it may be that raised background EC caused ally considered to be sensitive to disturbance, hence their by storage of salts in the landscape prevents recovery and world-wide use in the EPT biotic index (Lenat and Barbour recolonization during the unsalted season at the downstream 1994). However, this biotic index may not always be a sensi- Pipers Creek downstream site. If salinization is prevented, tive indicator of biological health. It is not unusual for some recolonization may be expected from upstream areas. When EPT taxa to be reduced, whereas taxa richness remains the systems do not recover, salt-sensitive biota are replaced by same, because they are being replaced by more salt tolerant more salt tolerant species (Hart et al. 2003). Disturbance species (Hart et al. 2003; Keford et al. 2010). EPT taxa can cause fragmentation of populations and recovery can be comprised a mean 47% (maximum 75%) of individual mac- slow, even in the absence of physicochemical barriers (Lake roinvertebrates found at all sites in this study. Even within et al. 2007). Ephemeroptera often recolonise downstream the EPT taxa there is a range of salt sensitivity. Increases in sites via drift but are known to be slow to recolonise streams abundances of more salt tolerant Trichoptera and Plecoptera through lateral movement because of limited fight abilities, may increase EPT taxa richness, which is not necessarily which often confne individuals to within ≈100 m of streams a good environmental outcome. For example, at the Pip- from which they originate (Petersen et al. 2004). ers Downstream site, Ephemeroptera were absent in 2017 Generally, more Oligochaeta were found downstream of samples, despite EPT families comprising 45–67% of total salted roads compared with their upstream site pairs in both taxa. Although found to be a good indicator of biological spring and autumn, with more observed in spring, particu- condition of macroinvertebrate assemblages in the United larly on Perisher Creek. However, by contrast Oligochaeta States (e.g., Cufney et al. 2010), lowland Australian streams were more abundant upstream compared with downstream (Chessman et al. 2006), and elsewhere (Lenat and Craw- at the Pipers Creek site-pair on three out of four sampling ford 1994; Lydy et al. 2000; Stark and Phillips 2009), this rounds, the exception being Spring 2016. This taxon may index does not appear to detect consistently the efects from be more sensitive to higher salinity than some other taxa

1 3 Archives of Environmental Contamination and Toxicology found in Australian alpine streams but more tolerant rela- Compliance with Ethical Standards tive to Ephemeroptera. This has been observed elsewhere in lowland Australian rivers (e.g., Keford et al. 2003). Oli- Conflict of interest The authors declare that they have no confict of gochaetes can be salt tolerant (e.g., Williams et al. 1999); interest. however, when Oligochaeta as a group are found to be toler- ant, it may be because the group includes particular species References with ecological traits that allow them to persist in conditions that are stressful to other species within the group. Also, ANZECC, ARMCANZ (2000) Australian and New Zealand Guide- when Oligochaetes are not identifed any further, it is impos- lines for freshwater and marine water quality. Volume 2, Aquatic sible to exclude changes to the mix of family, genera, or Ecosystems - Rationale and Background Information (Chapter 8). species of Oligochaetes present. The higher peak salinities Australian and New Zealand Environmental and Conservation and longer duration of salinity peaks observed during 2017 Council (ANZECC) and Agriculture and Resource Management Council of Australian and New Zealand (ARMCANZ), Canberra, support the idea that Australian alpine oligochaetes might be Australia. more sensitive to the high salinity inputs seen at the Pipers Baek MJ, Yoon TJ, Kim DG, Lee CY, Cho K, Bae YJ (2014) Efects downstream site than assumed, as do the initial RTT results of road deicer runof on benthic macroinvertebrate communi- of this study that were used to generate SSDs for Australian ties in Korean freshwaters with toxicity tests of calcium chloride (CaCl2). Water Air Soil Pollut 225:1961. https://doi.​ org/​ 10.​ 1007/​ ​ alpine taxa. s11270-​014-​1961-6 Our study observed that Australian alpine taxa in mac- Benke AC, Wallace JB (1980) Trophic basis of production among net- roinvertebrate communities are more sensitive to salin- spinning caddisfies in a southern Appalachian stream. Ecology ity inputs than those found in Australian lowland streams 61:108–118. https://​doi.​org/​10.​2307/​19371​61 Bray J, Reich J, Nichols S, Kon Kam King G, MacNally R, Thompson (Fig. 4). Estimates on the efect of salinity from the rela- R, O’Reilly-Nugent A, Keford B (2019) Biological interactions tively large existing dataset of Australian lowland mac- mediate context and species-specifc sensitivities to salinity. Philo- roinvertebrates will likely tend to underestimate efects in sophical Trans R Soc B 374:20180020. https://​doi.​org/​10.​1098/​ Australian alpine streams. Consequently, these data from rstb.​2018.​0020 Bray JR, Curtis JT (1957) An ordination of the upland forest communi- lowland streams should not be used in alpine streams. Fur- ties of southern Wisconsin. Ecol Monogr 27:325–349. https://doi.​ ​ thermore, this is at least in part because some taxonomic org/​10.​2307/​19422​68 groups common in natural alpine areas (i.e., Ephemeroptera Camp A, Funk D, Buchwalter D (2014) A stressful shortness of breath: in particular) are very sensitive to salinity relative to other molting disrupts breathing in the mayfy Cloeon dipterum. Fresh- water Sci 33:695–699. https://​doi.​org/​10.​1086/​677899 groups (Fig. 5). The taxa that were less abundant where Cañedo-Argüelles M, Hawkins CP, Keford BJ, Schäfer RB, Dyack B, de-icing salts were present tended to be the same taxa that Brucet S, Buchwalter D, Dunlop J, Frör O, Lazorchak J, Coring toxicity testing revealed to be relatively salt sensitive. This E, Fernandez HR, Goodfellow W, González Achem AL, Hatfeld- study fnds that de-icing salts are increasing stream salinity Dodds S, Karimov BK, Mensah P, Olson JR, Piscart C, Prat N, Ponsál S, Schulz CJ, Timpano AJ (2016) Saving freshwater from and altering macroinvertebrate community composition in salts. Science 351:914–916. https://doi.​ org/​ 10.​ 1126/​ scien​ ce.​ aad34​ ​ streams near Perisher that received run-of from road de- 88 icing activity. Cañedo Argüelles M, Keford BJ, Piscart C, Prat N, Schäfer RB, Schulz C-J (2013) Salinisation of rivers: an urgent ecological issue. Envi- ron Pollut 173:157–167. https://doi.​ org/​ 10.​ 1016/j.​ envpol.​ 2012.​ 10.​ ​ Supplementary Information The online version contains supplemen- 011 tary material available at https://doi.​ org/​ 10.​ 1007/​ s00244-​ 021-​ 00827-1​ . Charles S, Veber P, Delignette-Muller ML (2018) MOSAIC: a web-interface for statistical analyses in ecotoxicology. Envi- Acknowledgements The authors are grateful for project funding from ron Sci Pollut Res 25:11295–11302. https://​doi.​org/​10.​1007/​ the Australian Alps national parks Cooperative Management Pro- s11356-​017-​9809-4 gram, from which Mark Shenton and Benjamin Moulding received Chessman BC, Thurtell LA, Royal MJ (2006) Bioassessment in a harsh scholarships. This work was conducted under a Scientifc Licence (no. environment: a comparison of macroinvertebrate assemblages at SL101713) from the NSW Parks and Wildlife Service and Research reference and assessment sites in an Australian inland river sys- Permit (no. 10007909) from the Victorian Department of Environment, tem. Environ Monitor Assess 119:303–330. https://​doi.​org/​10.​ Land, Water and Planning. They also thank the staf of NSW Parks and 1007/​s10661-​005-​9027-2 Wildlife Service and Elizabeth Wright (Parks Victoria) for collection Clarke KL, Gorley RN (2006) Primer v6: User Manual/Tutorial. of samples and providing accommodation for researchers, as well as Primer-E Ltd, Plymouth Transport NSW contractors for providing road de-icing salt samples. Clements A, Suter P, Fussell M, Silvester E (2016) Macroinvertebrate Thanks to Simon Foster for assistance in IC analyses, Guillaume Kon communities in spring-fed alpine source pools. Hydrobiologia Kam King for advice and collaboration during the project, and Kathryn 777:119–138. https://​doi.​org/​10.​1007/​s10750-​016-​2770-2 McGilp for undertaking feld sampling and data collection. Addition- Clements WH, Kotalik C (2016) Efects of major ions on natural ben- ally, Kaylin de Lembracht, Brian Johnston, Callum McKinnon, Hannah thic communities: an experimental assessment of the US Environ- Moulding, Kylie Shenton, Mark Tupalski and Elizabeth Wright for mental Protection Agency aquatic life benchmark for conductivity. volunteer assistance in the feld. Finally, thanks to two anonymous Freshwater Sci 35:126–138. https://​doi.​org/​10.​1086/​685085 reviewers for their helpful comments that greatly improved the quality of this manuscript.

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