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Article

Effects of Cumulative Municipal Wastewater Exposure on Benthic Macroinvertebrate Assemblages: An Experimental Stream Approach

by
Aphra M. Sutherland
*,
Frederick J. Wrona
and
David C. Barrett
Department of Biological Sciences, University of Calgary, 2500 University Dr NW, Calgary, AB T2N 1N4, Canada
*
Author to whom correspondence should be addressed.
Hydrobiology 2025, 4(2), 17; https://doi.org/10.3390/hydrobiology4020017
Submission received: 13 February 2025 / Revised: 4 June 2025 / Accepted: 6 June 2025 / Published: 13 June 2025
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Abstract

Municipal wastewater effluent (MWWE) is a common source of nutrient enrichment and provides a route for emerging substances of concern (ESOCs) to enter aquatic systems. Community composition and abundance metrics of benthic macroinvertebrates are commonly utilized to assess ecological impacts associated with nutrient enrichment; however, the responses of these metrics in systems with diverse chemical mixtures from MWWE, are not well understood. This study specifically addresses the effects of cumulative loading of tertiary-treated MWWE through responses in benthic macroinvertebrate communities in experimental control and treatment streams. Treatment streams used source river water previously exposed to upstream wastewater treatment plants but with an additional 5% by volume tertiarily treated MWWE, while control streams used only source river water. Surbers and artificial substrate rock baskets were used to examine impacts on both established and colonizing benthic communities, respectively. No significant differences were observed between the control and treatment streams in any of the community metrics of well-established benthic communities. In contrast, significant decreases in colonizing taxon diversity and evenness were found between treatment and control streams. The dominant taxa (most abundant family, by percentage of sample) in the community, often filter feeders, significantly increased in percentage of the total community in treatment streams. This response was consistent with a nutrient enrichment effect, with no evidence of ESOC related toxicity. This study highlights the need for bioassessment programs to utilize approaches involving varied in-situ sampling methods and controlled exposure systems to gain a better understanding of how various stages of community-level development are impacted by urban pollutants such as MWWE.

1. Introduction

Municipal wastewater effluent (MWWE) is a major source of nutrients in waterways and a significant route for emerging substances of concern (ESOCs), including pharmaceuticals and personal care products, to enter aquatic systems [1,2,3,4]. The composition of MWWE is affected by changes in anthropogenic factors, such as human population size and chemical substance use, which can influence the effectiveness of wastewater treatment. This is particularly relevant for wastewater treatment plants (WWTPs) that are designed primarily to remove nutrients, pathogens, suspended solids, and keep chemical and biological oxygen demand within a certain range [2]. Therefore, MWWE may demonstrate high variation in loadings of nutrients and other chemical constituents including ESOCs over time and across treatment facilities [5]. Given this complexity, isolating specific impacts of MWWE constituents on aquatic ecosystem health is difficult, particularly in large lotic systems receiving multiple MWWE inputs and other point- and non-point source pollutants [6,7].
Biological endpoints are used in effects-based monitoring of complex pollutants to characterize impacts that could manifest differently through multiple possible pathways [8,9]. Benthic macroinvertebrate communities are a biological endpoint often used to assess impacts from aquatic effluents, including nutrient enrichment [10,11]. MWWE exposure is known to reduce diversity, richness, and the presence of organics-sensitive taxa in macroinvertebrate communities through the eutrophication of aquatic systems [4,12]. In contrast, the effects of ESOCs on macroinvertebrate communities are not well understood, as the physiological sensitivities and responses of taxa are associated with different outcomes for many pharmaceuticals and personal care products, ranging from toxicological impacts to no notable effects [13,14,15]. While the quality of the final MWWE is monitored to meet regulatory requirements prior to release, the assessment of subsequent ecological effects in the receiving environment is not necessarily performed. This produces potential to overlook the combined effects from effluent addition on aquatic ecosystems already impacted by upstream MWWE releases, which will further increase nutrient concentrations [16]. Additionally, while most ESOCs are not observed in high concentrations in final MWWE, multiple WWTP inputs could increase some ESOCs to biologically relevant and toxic concentrations [17]. Changes in MWWE loading can also be compounded by variation in flow conditions, which affects both macroinvertebrate communities and concentrations of contaminants [18,19]. While the cumulative impacts of MWWE in rivers, alongside pollutants such as pesticides, has been associated with significant changes in benthic macroinvertebrate communities, cumulative impacts from sequential WWTPs has typically focused on fish health [20,21]. The aggregated effects from sequential MWWE inputs remain poorly defined for benthic macroinvertebrates.
Assessing a common, diverse, and flux dependent pollutant such as MWWE through ecosystem responses should also require assessing community responses at different stages of development, multiple sampling methods, and ideally also controlled exposure systems, especially for refining our understanding of the pathways related to observed biological responses [11,19,22,23]. Experimental streams, which are effective at replicating the environmental conditions, have shown that both the composition and total concentration of MWWE have a significant impact on benthic macroinvertebrate assemblages [24,25]. Moreover, using a combination of sampling methods targeting different stages of community development (early colonizing versus well established communities) provides a more comprehensive understanding of how MWWE exposure affects aquatic ecosystems.
This study aimed to understand the effects of cumulative MWWE exposure on established and colonizing benthic community assemblages, in controlled experimental streams. Treatment streams were exposed to a 5% (by volume) addition of tertiary-treated MWWE, using source water from a nearby river that contained low background levels of MWWE from multiple upstream wastewater treatment plants. A combination of Surbers for established benthic communities and standardized rock baskets for colonizing communities, was used to assess the ability to distinguish between nutrient- and ESOC-associated effects. The findings provide new insights for improving the design and interpretation of in-situ effects-based MWWE biomonitoring programs.

2. Materials and Methods

2.1. The Advancing Canadian Water Assets Experimental Streams

The Advancing Canadian Water Assets (ACWA) experimental facility is associated with the Pine Creek tertiary wastewater treatment plant (WWTP) in Calgary, AB (50°51′43.2″ N 113°59′27.6″ W) and consists of replicated constructed streams that receive specified volumes of wastewater effluent at the headwater input of the stream (Figure 1). The outdoor streams allow streamside vegetation, benthic macroinvertebrates, autotrophic and heterotrophic bacteria and fungi, and fish to colonize naturally from the surrounding watershed, as well as through the direct intake of source water from the adjacent Bow River [23].
ACWA’s experimental stream facility hosts twelve ~320 m long replicated, and hydraulically isolated streams with identical alternating riffle (10 m) and pool (20 m) sequences. The streams are designed to match the substrate size composition of local 1st order streams. Temperature and dissolved oxygen regimes are generally consistent among the replicate control and treatment streams, though upstream to downstream gradient differences in pH are observed [23]. All streams are run as triplicates, including the source water only control streams and effluent treatment streams that receive the addition of 5% by volume tertiarily treated (with UV disinfection) MWWE.
Effects of nutrient enrichment correlated with MWWE in the source Bow River are observed downstream of three treatment plants in the city of Calgary, primarily through periphyton growth [26,27]. ESOCs related to MWWE inputs have been measured, with the highest concentrations associated with diabetes drugs, artificial sugars, antibiotics and analgesics [28]. Although ESOCs are present, no studies to date have examined their potential effects on basal components of the aquatic food web such as macroinvertebrate communities. The experimental streams have received Bow River source water and Pine Creek effluent in the same arrangement since their inception in 2010, resulting in long-established benthic invertebrate communities. Therefore, the control streams represent water quality conditions of source water downstream of two of Calgary’s three WWTPs, one of which is currently being upgraded to support over 1.3 million residents. All direct effluent treatments make up 5% of the total flow, which is continuously added through the water inflow (13.4 L/s).
This study undertook further characterization of the chemical and physical attributes of the replicate streams to confirm that the underlying abiotic conditions were comparable. ACWA measures head pond chemistry each month. The two months overlapping with rock basket deployment were used to characterize the composition of the source water in the streams. Water quality over the experimental period (September–October 2020) from the head pond feeding Bow River water into the experimental streams included temperature, pH, total organic carbon, total nitrogen, and total phosphorus. Water quality samples were taken by ACWA technical staff and processed in the attached water quality laboratory. Conductivity in the streams was monitored by in-situ loggers, measuring every 30 min.

2.2. Field and Laboratory Methods

Physical characteristics of the streams (average width, depth, substrate size) were assessed at both upstream (riffle 2) and downstream (riffle 10) sampling sites in each stream. Three measurements were taken at the top, middle, and bottom of the riffle and averaged. The substrate was characterized using an adaptation of the pebble count from the Substrate Characteristics section of the Canadian Aquatic Biomonitoring Network (CABiN) field manual for wadeable streams [29]. Fifty rocks were randomly chosen by taking the boot-end rock every 2 steps. Rocks were classified using Wentworth substrate classes of the length of the longest axis.
Benthic macroinvertebrate assemblages were quantified in the control and treatment streams using rock baskets to provide additional habitat control and identify colonization patterns, as well as more conventional (30.5 × 30.5 cm) and timed (30 s) Surber samples [30]. Rock baskets with standardized artificial substrate were placed in the experimental streams on 14 September 2020, and allowed to colonize for a period of six weeks [31]. The rock baskets were constructed of wire cages (30 × 24 × 5.5 cm) with a bottom layer of inert plastic scour pads covered with 40 identical ceramic briquettes (two rows of 20 briquettes arranged in a 5 × 4 pattern). To account for lower observed abundances and higher variability associated with rock baskets compared to samplers of established communities, three replicates were placed in both the upstream (approximately 20 m below effluent input) and downstream (approximately 300 m below effluent input) riffles in each replicate control and treatment stream [32,33]. Surber samples were taken on the day of rock basket removal from each replicate control and treatment experimental streams, one from upstream and one from downstream. The rock baskets were disassembled immediately beside the streams into scour pad and briquette bags. Associated biological material from the wire cage and captured while lifting rock baskets out of the substrate was frozen at −20 °C until processing. In total, 6 Surbers and 18 rock baskets were sampled for each treatment stream.
Both rock basket and Surber samples were subsampled across two fractions, using a 850 µm sieve to separate heavy macroinvertebrates, while the remainder was sampled using an Imhoff cone [34]. For both sampling methods, subsamples ranged from 25% to 100% of the total sample, depending on the overall abundance of taxa. A subsample was required to contain a minimum of 100 organisms for identification and was resampled until the sample either met that threshold or was 100% identified, to establish the distribution of major taxonomic groups (primarily order) [34,35]. Taxa were removed from debris under a dissecting microscope and identified to the lowest practical unit (predominantly family), following Aquatic Invertebrates of Alberta [36]. Samples were scaled by multiplying by the coefficient of subsampling fraction to standardize sorting effort before analysis.

2.3. Ecological Methods and Statistical Analysis

Using the vegan package, communities were characterized for macroinvertebrate assemblages for both sampling methods, using a range of ecological metrics [37]. Community endpoints included total raw abundance, richness, Shannon-Weiner diversity, Pielou’s evenness, the % contribution from the dominant taxa, and the Hilsenhoff Biotic Index (HBI) for the stream. All metrics, apart from abundance, were calculated at taxonomical resolution of family. These metrics were tested for equal variance, prior to comparison using permutation tests. Differences between distribution of taxa into different taxonomic orders were assessed using a two-sided Kolmogorov-Smirnov test, which tested whether the major taxonomic groups in control and treatment streams had significantly different cumulative frequency distributions [38].
Prior to multivariate analysis, benthic macroinvertebrate data was log10 transformed to reduce the influence of extreme values in the data set, in this case primarily zeroes [39]. Counts were also multiplied by a factor of 0.2 to down-weight the contribution of rare taxa [40]. A follow-up Similarity Percentages (SIMPER) analysis determined which taxa were most responsible for differences between sites using the simper function from the vegan package [11,37]. Bray-Curtis distances were used to perform a Principal Coordinate Analysis (PCoA), which spatially represented replicate similarity between sampling sites in streams for both sample types [37]. All analysis was performed using RStudio (R version 4.3.2).

3. Results

3.1. Physicochemical Parameters of Artificial Streams

All streams showed comparable width, depths, and mean substrate size and all parameters were relatively stable between the two monthly measurements (Table 1). Two exceptions include total phosphorus, which dropped by ~50% from the start to end of the rock basket exposure, and temperature, which also almost halved during the study. Given the similarity in stream conditions, all samples taken from replicate streams were pooled into control and treatment groups.

3.2. Diversity Metrics

Most community metrics were similar between control and treatment streams for different community stages (Table S1). The total raw abundance of all combined macroinvertebrates identified in samples was similar between treatments. The mean family richness was almost identical for all communities (~9 families) with a total of 22 unique families observed in the established communities and 28 unique families observed in the colonizing communities. Additionally, neither abundance nor richness were significantly different between treatments (Figure 2). At both community stages, family-level HBI did not significantly differ between control and treatment streams. Generally, each experimental stream had healthy ratings at “Fair”, with one colonizing replicate at “Good” and 36% of the colonizing replicates and 27% of the established replicates at “Fairly Poor” (Figure 3). In established benthic communities, the mean diversity, evenness, and the relative percentage of the dominant taxa also did not significantly differ. However, in the colonizing communities, the Shannon-Weiner diversity index was higher in control streams (mean ± SD = 1.35 ± 0.37) than in treatment streams (mean ± SD = 1.11 ± 0.30), and the mean Pielou’s evenness index was also higher in control streams (mean ± SD = 0.61 ± 0.16) compared to the treatment stream (mean ± SD = 0.50 ± 0.11). In contrast, the mean relative percent of the dominant taxa was higher in treatment streams (mean ± SD = 60.81% ± 0.30) than control stream (mean ± SD = 45.78% ± 0.37%) (Table S2).

3.3. Major Taxonomic Group Distribution

The frequencies of the abundance of each taxonomic order or other major taxonomic groups represented in colonizing communities were not significantly different between control and treatment streams (two-Sample Kolmogorov-Smirnov test, D = 0.4, p = 0.873) (Figure 4) (Table S3). Diptera or Ephemeroptera, Plecoptera, and Trichoptera (EPT) were always the most dominant taxa in both streams. However, the group EPT was generally represented by only Hydropsychidae, a Trichoptera family particularly tolerant of organic pollution. Diptera were primarily represented by the families Simuliidae and Chironomidae, which were common in both treatments across both community stages but were more dominant in colonizing communities, by relative percentage of the community. The next most observed group was the order Coleoptera, but in much smaller numbers. Gastropoda and Amphipoda were also represented by limited representatives. Gastropods, Acari, and Collembola were only found in colonizing benthic communities.
Major taxonomic groups in the established communities also did not have significantly different cumulative frequency distributions of major taxonomic groups between stream types (two-Sample Kolmogorov-Smirnov test, D = 0.2, p = 1) (Table S3). Similarly, the dominant groups in both streams were representatives of Diptera and EPT orders. The most common taxa in the treatment stream were from the EPT group, which was dominated by the same Trichoptera family (Hydropsychidae) as the rock basket samples. Diptera taxa, proportionally, represented about twice as much of the community in control streams. Otherwise, the streams were similar, with Amphipoda and Coleoptera having consistent presence, as well as a small number of leeches (Hirudinea). Hirudinea were only found in established communities.

3.4. Multivariate Comparisons

The SIMPER analysis between control and treatment streams identified that Simuliidae contributed to about ~30% of Bray-Curtis differences in colonizing samples and ~20% in established samples. Simuliidae were present in both streams, but the control stream accounted for about 7% more of the total number observed in colonizing communities. Colonizing communities had ~40% higher numbers of Chironomidae in the control streams, which also contributed to the difference between stream types. Between two Diptera families, 75% of the difference between colonizing communities in different treatments was explained. Established communities were also highly differentiated by Hydropsychidae (~30%), Hyallelidae (~15%), and Elmidae (~15%). All together, these families explained 78.6% of the Bray-Curtis differences between established communities in the streams. Control streams had higher proportions of Simuliidae, while treatment streams generally had more Hydropsychidae, Hyallelidae, and Elmidae. Other taxa contributed to less than 5% of the difference between streams (Figure 5).
A PCoA showed that control and treatment streams tended to be similar for both colonizing and established communities (Figure 6). Two components, for colonizing communities, cumulatively explained ~25% of the variation between sites (MSD1 = 15%, MSD2 = 10%) (Table S4). In the established samples, two components cumulatively explained 62% of the variation between sites (MSD1 = 45%, MSD2 16%) (Table S4). The overlap between streams in the established samples showed strong similarities between control and treatment communities. The colonizing communities also displayed consistent overlap between control and treatment streams but with more room for variation between the streams.

4. Discussion

Overall, there was limited evidence that a 5% addition of MWWE to a baseline exposure resulted in significant change to the benthic macroinvertebrate communities, as represented through different developmental stages of benthic macroinvertebrate community structure and abundance patterns. Established communities for benthic macroinvertebrates revealed no differences in community metrics or cumulative frequency distribution of major taxonomic groups and overall similarity between control and treatment streams was high. In contrast, significant differences were observed in the colonizing communities between diversity, evenness, and the relative percentage of the dominant taxa in the community. However, macroinvertebrate abundance, richness, and HBI between the control and treatment streams demonstrated no difference, and neither did the cumulative frequencies of major taxonomic groups. This lack of differences indicates that there is no evidence for significant toxicological effects from increased nutrient levels or ESOCs from MWWE and that the cumulative exposure to MWWE at these levels is indicative of nutrient enrichment [6]. However, as the sampling replication of the colonized substrate was triple that of the established communities to account for lower abundances and richness, as well as higher variation, observed in other rock basket or artificial versus natural substrate studies, a higher replication of samples in established communities could observe similar effects to those identified in the colonizing communities ties [30,32,33].
Other mesocosm experiments involving exposure to MWWE have not previously addressed cumulative exposure, but a similar 5% addition of MWWE, compared to a control in a different stream mesocosm study, did result in change to established benthic macroinvertebrate assemblages [41]. In the source river (the Bow River), in-situ monitoring showed that benthic macroinvertebrate assemblages downstream of multiple point source MWWE sources have increased macroinvertebrate abundances, lower diversity and richness, and an increase in pollution tolerant taxa [42,43]. However, there was no corresponding decline in abundance and richness or a rise in dominance that would indicate strong toxic effects from eutrophication or high ESOC concentrations [24]. In the current study of the ACWA streams, the observed patterns between established communities did not fully match these previously observed differences. Rather, the results of this study suggest that a cumulative addition of 5% of the same MWWE to an already nutrient enriched system may not significantly change benthic macroinvertebrate communities, at least to the same extent as a similar non-cumulative input might [26,41,43]. This helps to understand of how benthic macroinvertebrate communities may respond to MWWE exposure in systems with multiple inputs, particularly in the northern mid-latitudes. While benthic communities in the source river have not been investigated for direct latitudinal effects, food sources, such as periphyton, are impacted by additional limitations through changes to light availability. A study in the source river indicated that while phosphorus availability stimulated growth of periphyton in the source river for the artificial streams, growth was also limited by light, river temperature, and discharge [44]. This is relevant to understanding how cumulative nutrient effects may manifest in this geographic area.
Colonizing benthic macroinvertebrate communities were significantly different when exposed to a cumulative 5% input. In systems where hydrological patterns regularly alter and remove habitat, understanding how colonization patterns are impacted by MWWE provides new insights into how resilient re-colonizing communities are to increased effluent loading [45,46]. Concentrations of nutrients and ESOCs typically increase during low-flow periods, as these are highly linked to MWWE releases, as treated effluent volume does not correspond to hydrological shifts [18,28]. This is particularly of interest in aquatic systems near growing population centres that experience regular and significant changes in flow, such as glacier-fed or flow-regulated rivers. However, while previous studies indicate that a six-week deployment produces a stabilized community, this may also depend on the environment where samplers are deployed [47]. These rock basket communities did not closely resemble established communities, which may have meant they had not fully stabilized; however, rock basket samplers do not always resemble communities measured in more conventional sampling methods, as they will support different, and generally fewer, functional lifestyles [30]. Particularly, the microhabitats formed by the identical substrates may support different and less diverse communities [48].
The specific patterns observed in rock baskets do suggest the presence of a biological response in colonizing communities. The most dominant taxa observed in the artificial substrate samplers were detritus feeders (e.g., Chironomidae, Simuliidae, and Hydropsychidae), which were also present in the established communities, but were less uniformly dominant [4]. It is possible that differences could be related to increased depositional sediment from nutrient induced growth in the system, as increased organic material is associated with MWWE. This accumulation would likely be a key dissimilarity between established and colonizing samplers, as established communities would have more diverse food sources for benthic macroinvertebrates. In similar colonization experiments of benthic macroinvertebrates into artificial substrate samplers, detritus feeders are often the initial colonizers [49,50]. The most successful taxa colonizing clean substrate is organisms that are small, tolerant, and have very short lifecycles, which has been observed to be as short as 10 days for Chironomids [51,52]. The metrics with significant differences (diversity, evenness, and% dominance) in the colonizing communities is a preliminary indication that these organisms can take advantage of uncolonized substrate more effectively than other organisms, particularly if MWWE may be increasing the availability of organic material as a food source. This is important to track in rivers where flow can dramatically change, and colonization processes may prevail in significant portions of available habitat. Changes in colonization processes from cumulative effects of MWWE could have long-term impacts on resilience of freshwater systems [46]. For instance, some taxa could demonstrate patchier distributions, due to fragmentation of habitat [53]. While lotic communities may restabilize post-disturbance, this relies on colonization patterns remaining predictable.
In-situ monitoring of aquatic ecosystems is necessary, given the high likelihood of increasing MWWE loadings from WWTPs. However, it can be difficult to isolate and understand specific mechanisms for ecological impacts from complex effluents, such as MWWE [54]. This project’s use of experimental stream approach provided a more controlled but naturalized system in which to identify how cumulative inputs of MWWE affected benthic macroinvertebrate communities. Experimental stream approaches can be used to better predict effects in systems with multiple WWTPs. Including both established and colonizing communities in this assessment revealed that colonizing communities are dominated by different taxa, primarily filter feeders tolerant of organic pollution, in the presence of cumulative MWWE inputs. Established communities did not show a significant response to cumulative MWWE inputs, perhaps related to lower sampling replication. However, the response in colonizing communities is a reminder that different community stages may vary in sensitivity to contaminants and biomonitoring programs should make efforts to assess multiple stages to identify what stages might provide earlier warnings for impacts.

5. Conclusions

Experimental streams located at the ACWA facility showed that cumulative impacts of MWWE affected the diversity and evenness of benthic macroinvertebrate colonizing communities. Streams exposed to additional MWWE inputs (5% added to a stream containing background levels of MWWE from upstream WWTPs in the river feeding the streams) were more susceptible to being dominated by early colonizers, generally filter feeders, possibly signaling an increase in accumulated sediment. Many early colonizers were pollution tolerant organisms, which could change how benthic macroinvertebrate communities re-establish if increases in MWWE exposure exacerbated this pattern too far. This is particularly concerning in rivers with large variability in flow conditions where colonization is are a regular feature of communities as water provides or removes habitat. In contrast, established benthic macroinvertebrate communities did not demonstrate a response to the cumulative exposure to MWWE under experimental conditions. Overall, differences in community endpoints generally align with non-chemical impacts of MWWE, such as increased depositional sediment, and do not provide evidence of impact from ESOCs through evidence of toxicological effects. Patterns perceived in these more highly controlled mesocosm experiments across multiple stages of community development should be considered when establishing biological indicators and sampling methods for assessments of cumulative MWWE impacts, particularly as populations grow and components of MWWE evolve.

Supplementary Materials

The following supporting information can be downloaded at: https://www.mdpi.com/article/10.3390/hydrobiology4020017/s1, Table S1: Community metrics comparing benthic macroinvertebrate abundance and community using artificial substrate rock baskets samplers (n = 18) deployed in the streams for six weeks in the fall of 2020 and conventional grab (Surber) samples (n = 6). Table S2: Permutation test results (100,000 resamples,) of community metric comparisons between control (background municipal wastewater effluent signal) and treatment (5% municipal wastewater effluent signal added) streams at the Advancing Canadian Water Assets experimental stream facility in Calgary, AB. Streams were sampled using artificial substrate rock baskets samplers (n = 18 per treatment) deployed in the streams for six weeks in the fall of 2020 and conventional grab (Surber) samples (n = 6 per treatment). * Signifies significance at p < 0.05. Table S3: Results of a 2-way Kolmogorov-Smirnov test comparing frequency distributions of major taxonomic groups in using artificial substrate rock baskets samplers (n = 18) deployed in the streams for six weeks in the fall of 2020 and conventional grab (Surber) samples (n = 6). Table S4: Results of the PCoA and variance explained by top two components. Benthic macroinvertebrates were sampled using artificial substrate rock baskets samplers (n = 18) deployed in the streams for six weeks in the fall of 2020 and conventional grab (Surber) samples (n = 6). Figure S1: Reference photo of constructed rock basket removed from ACWA streams after six weeks (September to October 2020).

Author Contributions

Conceptualization, A.M.S., F.J.W. and D.C.B.; Methodology, A.M.S., F.J.W. and D.C.B.; Software, A.M.S.; Validation, A.M.S., F.J.W. and D.C.B.; Formal Analysis, A.M.S. Investigation, A.M.S., F.J.W. and D.C.B.; Resources, F.J.W.; Data Curation, A.M.S.; Writing—Original Draft Preparation, A.M.S.; Writing—Review & Editing, A.M.S., F.J.W. and D.C.B.; Visualization, A.M.S., F.J.W. and D.C.B.; Supervision, F.J.W.; Project Administration, F.J.W. and D.C.B.; Funding Acquisition, F.J.W. and D.C.B. All authors have read and agreed to the published version of the manuscript.

Funding

This research was funded by the City of Calgary, Natural Sciences and Engineering Research Council Discovery Alliance Grant ALLRP 567652-21, Natural Sciences and Engineering Research Council Discovery Grant RGPIN/05146-2018, Alberta Conservation Association Research grant 020-00-90-303, and SVARE Endowed Research Chair (FJW).

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Data Availability Statement

Dataset available on request from the authors.

Acknowledgments

The authors acknowledge water chemistry data shared by the Advancing Canadian Water Assets (ACWA) facility. The authors extend thanks for all assistance provided by C. O’Grady, N. Ruecker, and other supporting staff from the City of Calgary. Additional field support was provided by K.M., B.S., K.P., J.Z., S.M., and A.L. This work took place as part of the University of Calgary-City of Calgary Urban Alliance research collaboration initiative.

Conflicts of Interest

The authors declare no conflicts of interest.

Abbreviations

The following abbreviations are used in this manuscript:
MWWEMunicipal Wastewater Effluent
ESOCEmerging Substance of Concern
WWTPWastewater Treatment Plant
ACWAAdvancing Canadian Water Assets
CABiNCanadian Aquatic Biomonitoring Network
HBIHilsenhoff Biotic Index
SIMPERSimilarity Percentages
PCoAPrincipal Coordinates Analysis
EPTEphemeroptera Plecoptera Trichoptera

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Figure 1. A schematic of the experimental stream facility at Advancing Canadian Water Assets facility. For each sample site (stars), three rock baskets were deployed and a Surber sample taken. Control streams use Bow River water exposed to upstream wastewater treatment plant effluents. Treatment streams also use Bow River water but 5% of the total stream volume is replaced with tertiarily treated wastewater effluent. Streams using treatments not studied have their inflows marked in grey.
Figure 1. A schematic of the experimental stream facility at Advancing Canadian Water Assets facility. For each sample site (stars), three rock baskets were deployed and a Surber sample taken. Control streams use Bow River water exposed to upstream wastewater treatment plant effluents. Treatment streams also use Bow River water but 5% of the total stream volume is replaced with tertiarily treated wastewater effluent. Streams using treatments not studied have their inflows marked in grey.
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Figure 2. (A): abundance (count of total individuals), (B): richness (count of unique families), (C): Shannon-Weiner diversity, and (D): Pielou’s evenness characterizing benthic macroinvertebrate communities in control (background municipal wastewater effluent signal) and treatment (5% municipal wastewater effluent added) streams at the Advancing Canadian Water Assets experimental stream facility in Calgary. Streams were sampled using artificial substrate rock baskets for colonizing benthic communities (n = 18 per treatment) deployed for six weeks in the fall of 2020 and conventional grab (Surber) samples for established benthic communities (n = 6 per treatment).
Figure 2. (A): abundance (count of total individuals), (B): richness (count of unique families), (C): Shannon-Weiner diversity, and (D): Pielou’s evenness characterizing benthic macroinvertebrate communities in control (background municipal wastewater effluent signal) and treatment (5% municipal wastewater effluent added) streams at the Advancing Canadian Water Assets experimental stream facility in Calgary. Streams were sampled using artificial substrate rock baskets for colonizing benthic communities (n = 18 per treatment) deployed for six weeks in the fall of 2020 and conventional grab (Surber) samples for established benthic communities (n = 6 per treatment).
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Figure 3. (A): the Hilsenhoff Biotic Index (HBI)) characterizing benthic macroinvertebrate communities and, (B): relative percent contribution of the dominant taxa in control (background municipal wastewater effluent signal) and treatment (5% municipal wastewater effluent added) streams at the Advancing Canadian Water Assets experimental stream facility in Calgary. Streams were sampled using artificial substrate rock baskets for colonizing benthic communities (n = 18 per treatment) deployed for six weeks in the fall of 2020 and conventional grab (Surber) samples for established benthic communities (n = 6 per treatment).
Figure 3. (A): the Hilsenhoff Biotic Index (HBI)) characterizing benthic macroinvertebrate communities and, (B): relative percent contribution of the dominant taxa in control (background municipal wastewater effluent signal) and treatment (5% municipal wastewater effluent added) streams at the Advancing Canadian Water Assets experimental stream facility in Calgary. Streams were sampled using artificial substrate rock baskets for colonizing benthic communities (n = 18 per treatment) deployed for six weeks in the fall of 2020 and conventional grab (Surber) samples for established benthic communities (n = 6 per treatment).
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Figure 4. The proportion of high-level taxonomic groups (mostly Order) in control (background municipal wastewater effluent signal) and treatment (5% municipal wastewater effluent added) streams at the Advancing Canadian Water Assets experimental stream facility in Calgary. Streams were sampled using (A) artificial substrate rock baskets for colonizing benthic communities (n = 18 per treatment) deployed for six weeks in the fall of 2020 and (B) conventional grab (Surber) samples for established benthic communities (n = 6 per treatment).
Figure 4. The proportion of high-level taxonomic groups (mostly Order) in control (background municipal wastewater effluent signal) and treatment (5% municipal wastewater effluent added) streams at the Advancing Canadian Water Assets experimental stream facility in Calgary. Streams were sampled using (A) artificial substrate rock baskets for colonizing benthic communities (n = 18 per treatment) deployed for six weeks in the fall of 2020 and (B) conventional grab (Surber) samples for established benthic communities (n = 6 per treatment).
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Figure 5. SIMPER analysis results from benthic macroinvertebrate samples taken from control (background municipal wastewater effluent signal) and treatment (5% municipal wastewater effluent added) streams at the Advancing Canadian Water Assets experimental stream facility in Calgary. Taxa contributing to over 5% of the difference (Bray-Curtis) between streams are shown in order of their contribution to the total cumulative difference for each method. Streams were sampled using artificial substrate rock baskets for colonizing benthic communities (n = 18 per treatment) deployed for six weeks in the fall of 2020 and conventional grab (Surber) samples for established benthic communities (n = 6 per treatment).
Figure 5. SIMPER analysis results from benthic macroinvertebrate samples taken from control (background municipal wastewater effluent signal) and treatment (5% municipal wastewater effluent added) streams at the Advancing Canadian Water Assets experimental stream facility in Calgary. Taxa contributing to over 5% of the difference (Bray-Curtis) between streams are shown in order of their contribution to the total cumulative difference for each method. Streams were sampled using artificial substrate rock baskets for colonizing benthic communities (n = 18 per treatment) deployed for six weeks in the fall of 2020 and conventional grab (Surber) samples for established benthic communities (n = 6 per treatment).
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Figure 6. PCoA analysis results from benthic macroinvertebrate samples taken from control (background municipal wastewater effluent signal) and treatment (5% municipal wastewater effluent added) streams at the Advancing Canadian Water Assets experimental stream facility in Calgary. The variance explained by each axis is labelled. Streams were sampled using (A) artificial substrate rock baskets for colonizing benthic communities (n = 18 per treatment) deployed for six weeks in the fall of 2020 and (B) conventional grab (Surber) samples for established benthic communities (n = 6 per treatment).
Figure 6. PCoA analysis results from benthic macroinvertebrate samples taken from control (background municipal wastewater effluent signal) and treatment (5% municipal wastewater effluent added) streams at the Advancing Canadian Water Assets experimental stream facility in Calgary. The variance explained by each axis is labelled. Streams were sampled using (A) artificial substrate rock baskets for colonizing benthic communities (n = 18 per treatment) deployed for six weeks in the fall of 2020 and (B) conventional grab (Surber) samples for established benthic communities (n = 6 per treatment).
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Table 1. Physical and chemical parameters in the control streams and treatment streams, including the water chemistry of the Bow River source head pond during the experimental period (September–October 2020). Physical attributes (flow, width, depth) of each stream are based on a stream average, calculated average from three measurements in two riffles (upstream and downstream). Mean substrate size was determined by taking 50 random stone samples at each riffle site. Mean conductivity is measured from in-situ loggers, measuring conductivity every 30 min, deployed in upstream and downstream locations in selected streams over the experimental period (control stream n = 3 and treatment stream n = 4).
Table 1. Physical and chemical parameters in the control streams and treatment streams, including the water chemistry of the Bow River source head pond during the experimental period (September–October 2020). Physical attributes (flow, width, depth) of each stream are based on a stream average, calculated average from three measurements in two riffles (upstream and downstream). Mean substrate size was determined by taking 50 random stone samples at each riffle site. Mean conductivity is measured from in-situ loggers, measuring conductivity every 30 min, deployed in upstream and downstream locations in selected streams over the experimental period (control stream n = 3 and treatment stream n = 4).
SiteTemperature (°C)pHTotal Organic Carbon (mg/L)Total Nitrogen
(mg/L)		Total Phosphorus
(mg/L)
Head pond
(September)		13.758.031.650.850.03
Head pond
(October)		7.558.191.40.750.015
SiteFlow
(L/s)		Mean Width (cm)Mean Depth (cm)Mean Substrate Size (mm)Mean Specific
Conductivity (µS/cm)
Control Streams13.4143.538.2860.6299.4
Treatment Streams13.4145.108.0559.0286.0
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Sutherland, A.M.; Wrona, F.J.; Barrett, D.C. Effects of Cumulative Municipal Wastewater Exposure on Benthic Macroinvertebrate Assemblages: An Experimental Stream Approach. Hydrobiology 2025, 4, 17. https://doi.org/10.3390/hydrobiology4020017

AMA Style

Sutherland AM, Wrona FJ, Barrett DC. Effects of Cumulative Municipal Wastewater Exposure on Benthic Macroinvertebrate Assemblages: An Experimental Stream Approach. Hydrobiology. 2025; 4(2):17. https://doi.org/10.3390/hydrobiology4020017

Chicago/Turabian Style

Sutherland, Aphra M., Frederick J. Wrona, and David C. Barrett. 2025. "Effects of Cumulative Municipal Wastewater Exposure on Benthic Macroinvertebrate Assemblages: An Experimental Stream Approach" Hydrobiology 4, no. 2: 17. https://doi.org/10.3390/hydrobiology4020017

APA Style

Sutherland, A. M., Wrona, F. J., & Barrett, D. C. (2025). Effects of Cumulative Municipal Wastewater Exposure on Benthic Macroinvertebrate Assemblages: An Experimental Stream Approach. Hydrobiology, 4(2), 17. https://doi.org/10.3390/hydrobiology4020017

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