Benthic Invertebrate Community Responses Associated with Mixed Runoff Contamination in Irrigation Canals: A Standardized Colonization-Plate Approach

2.1 Study design and reach selection

A comparative field design was employed to evaluate the exposure to contaminants and colonization by benthic macroinvertebrates in irrigation canals. Reaches were rated as either reference or impacted following the reconnaissance of land use, visual discharge points, bank condition, sediment deposition, and surrounding agricultural and peri-urban activities. Reference reaches were located upstream of large drainage inlets and were identified as having less obvious effects of influence from the drainage and more stable riparian margins. Impacted reaches were located downstream of mixed agricultural return flow and peri-urban runoff inputs and were identified with visibly increased fine sediment deposition and turbidity. Reaches were chosen to be as similar as possible in channel characteristics, channel depth, flow velocity, and substrate texture to minimize the interpretation of biological differences in response to gross differences in channel characteristics.

2.2 Study area description

The study was conducted in an agricultural–peri-urban irrigation-canal network in central Iraq, within a WGS84 coordinate corridor of 33.277–33.295°N and 44.260–44.291°E. The spatial design comprised 10 independent canal reaches/segments: five upstream reference reaches (R1–R5) and five downstream impacted reaches (I1–I5), with five Hester–Dendy plates per reach (n = 25 plates per exposure group; n = 50 total). Reference reaches were located 0.25–2.05 km from the upstream boundary, whereas impacted reaches were located 3.30–5.50 km downstream of the main transition/runoff-entry zone. The reference-to-impacted separation ranged from approximately 1.25 to 5.25 km along the canal corridor. Agricultural-return and peri-urban drainage inlets were located approximately 40–100 m from impacted reaches and more than 800 m from reference reaches. Figure 1 presents the georeferenced study-area map, including coordinates, reference and impacted reaches, runoff inputs, plate locations, and surrounding agricultural/peri-urban land use.

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Figure 1. Georeferenced study-area map showing reference and impacted irrigation-canal reaches, runoff inputs, Hester–Dendy plate locations, and surrounding agricultural and peri-urban land-use zones. Coordinates are reported in the WGS84 reference system

2.3 Colonization plates and field deployment

The experimental unit was the individual Hester–Dendy sampler. Fifty samplers were deployed for 28 days during a stable irrigation period: 25 in reference reaches, and 25 in impacted reaches. Each sampler comprised a series of hardboard plates separated by spacer blocks to provide repeated interstitial colonization surfaces. Before deployment, samplers were rinsed with deionized water, handled with powder-free gloves, labeled, and stored in clean containers. Within each reach, samplers were deployed in a blocked arrangement to minimize small-scale variation in depth, turbulence, and lateral position. Plates were suspended from permanent anchors at approximately mid-water depth, away from banks, macrophyte mats, eddy, and point inflows. At retrieval, each sampler was enclosed in a fine net to minimize loss of organisms and loose biofilm, and then transferred to a labeled container containing site water for laboratory processing. All 50 deployed samplers were recovered and retained for analysis, and comparisons were standardized by exposure group and the 28-day deployment period.

2.4 Water quality, microplastics, and trace metals

A calibrated portable multiparameter meter was used to measure in situ the temperature, pH, specific conductivity, dissolved oxygen, turbidity, and depth at each colonization site. Water-quality and hydraulic parameters were measured according to the field sampling schedule established for the colonization experiment. Measurements were conducted at deployment and retrieval of colonization plates (Day 0 and Day 28, respectively) at comparable times of day to minimize diel variation. Where repeated measurements were available, mean values across the deployment period were calculated for comparison between reference and impacted reaches. Because the study represents a single colonization period, temporal fluctuations associated with rainfall events, irrigation intensity, and seasonal changes may not be fully captured. Timed floats over a fixed distance were used to estimate the surface velocity, which was then used to calculate the approximate mean column velocity based on a consistent hydraulic coefficient, as per the typical assumptions associated with rough-bed flows [8]. The plate-associated biofilm was removed from the plates using gentle brushing and rinsing with filtered deionized water in clean stainless-steel equipment for the microplastic analysis. Surficial fine sediments were sampled within approximately 1 m of each plate, dried, homogenized, and subsampled by dry mass. A density separation was carried out using a zinc chloride solution, and then filtration was performed, and polymers were confirmed by pyrolysis-gas chromatography-mass spectrometry. Microplastics were reported as items kg⁻¹ dry sediment and µg g⁻¹ dry biofilm. The samples were filtered through 0.45-micrometer membranes and acidified using ultrapure nitric acid for the analysis of zinc and copper in the water. The fine sediments were sieved to the less than 63 micrometer fraction and subjected to microwave-assisted acid digestion. All digests, dissolved samples, and sediment were analyzed by inductively coupled plasma optical emission spectrometry, and all sediment concentrations were calculated on a dry-weight basis. Microplastic analysis followed contamination-control procedures, including procedural blanks, field blanks, and laboratory air-exposure controls. Samples were filtered on clean analytical membranes before polymer confirmation by pyrolysis-gas chromatography-mass spectrometry. For the comparative analytical framework, the microplastic workflow used a standardized protocol based on 0.45-µm filtration before polymer confirmation; therefore, the microplastic results are interpreted as comparative sediment and biofilm loads rather than a full size-resolved particle inventory. Recovery experiments were conducted using known quantities of reference polymer particles, and percentage recovery was calculated for each analytical batch. In order to identify a polymer, the pyrolysis-gas chromatography-mass spectrometry spectra had to match a reference polymer library with reasonable confidence. Where necessary, blank corrections were applied, and results were reported after the background contamination had been adjusted.

2.5 Macroinvertebrate processing and biotic indices

All samplers were disassembled in the laboratory on a clean tray and rinsed over a 500-micrometer sieve. Organisms retained and associated detritus were preserved and sorted under a stereomicroscope to the lowest possible taxonomic level, typically to the family or genus level, depending on the diagnostic characters and condition of the preserved material. To preserve sensitivity to the rare and pollution-sensitive taxa, whole samples were processed instead of subsampled. The taxon-by-sample abundance data were subsequently prepared, and total taxon richness, Shannon diversity, EPT richness, percentage EPT, BMWP, and ASPT were calculated. Two well-known family-level biotic indices of pollution-sensitive macroinvertebrate assemblages, BMWP and ASPT, were used. Based on the taxonomic data, functional feeding groups were determined and classified as relative abundance categories: scrapers, shredders, collectors, and predators.

2.6 Identification of macroinvertebrates

Macroinvertebrate identification was conducted to the lowest practicable taxonomic level, generally family level and, where diagnostic characters and preservation condition allowed, genus level, using standard freshwater macroinvertebrate keys. EPT taxa were assigned according to Ephemeroptera, Plecoptera, and Trichoptera membership. BMWP and ASPT were calculated as family-level comparative biotic indices. Because BMWP and ASPT were originally developed mainly for temperate European freshwater systems and have not been fully calibrated for Iraqi irrigation canals, their application was interpreted cautiously as a comparative indication of community response between reference and impacted reaches rather than as an absolute ecological-status classification.

2.7 Quality assurance

Quality assurance and quality control were documented in supplementary tables linked to the raw-data workbook. For microplastics, recovery tests were conducted for PE, PP, PET, PS, and PVC reference particles (n = 5 batches per polymer), together with procedural blanks, blank-correction procedures, limit of detection (LOD)/limit of quantification (LOQ) values, particle-size classes, morphology categories, and pyrolysis GC-MS library-match criteria. For metal analysis, the documented method included a 0.45-µm dissolved fraction, microwave-assisted digestion of the <63 µm sediment fraction, inductively coupled plasma optical emission spectrometry (ICP-OES) measurement, wavelength settings, detection limits, duplicate precision, and certified reference material (CRM)/spike-recovery benchmarks. All quality assurance/quality control (QA/QC) values are provided in the supplementary tables and individualized Excel sheets.

2.8 Statistical analysis

Data were entered into a validated spreadsheet containing the complete plate-level raw matrix for n = 25 plates per exposure group. To address potential pseudoreplication, the analysis incorporated 10 independent reaches/segments (five reference and five impacted), with five plates nested within each reach. Plate-level Welch t-tests were retained as descriptive comparisons, whereas the primary sensitivity analysis was repeated at the reach/segment level by aggregating plate values within each reach and comparing five reference-reach means with five impacted-reach means. The nested results are reported as reach-level sensitivity analysis and provided in the supplementary Excel workbook. Composite pollution index (CPI) was calculated from measured sediment microplastics, sediment Zn, and sediment Cu after z-score standardization. R version 4.3.x was used for statistical analysis; the accompanying workbook documents all values, formulas, and data labels.

CPI = z(sediment microplastics) + z(sediment Zn) + z(sediment Cu)

The CPI was not interpreted as a validated ecological risk index but rather as a statistical descriptor of the combined contamination gradient observed among sampling locations.

2.9 Data verification and supplementary documentation

The complete analytical dataset included group summaries and plate-level observations for the main physicochemical, contaminant, and macroinvertebrate outcomes, together with GPS coordinates, reach-level nesting identifiers, QA/QC records, polymer-characterization details, and validation matrices. Data-integrity checks included verification of sample identifiers, group assignments, analytical units, calculation formulas, summary statistics, and consistency between plate-level and reach-aggregated analyses. The accompanying Excel workbook provides the individualized raw data, reach-level summaries, QA/QC records, and sensitivity analyses used in the manuscript.

The most important finding of this study is that impacted reaches of the canal were hydrologically similar to the reference reaches but chemically and biologically degraded. Comparable temperature, depth, and velocity decrease the probability that the observed biological impairment was due to the differences in basic physical habitat. Conductivity, turbidity, nitrate, and phosphate were, however, quite high, and dissolved oxygen was low, reflecting a strong water-quality gradient related to runoff. Similar degradation has been reported in canals and surface water bodies that receive textile, tannery, industrial, municipal, or agricultural inputs [9]. The current trend is also in line with the general findings that under the pressures of pollutants, the oxygen and salinity conditions of canals and rivers are significantly affected by discharge management, dilution, and re-aeration capacity [10, 11]. Where effluent treatment is weak and particulate loads are high, suspended solids, nutrients, and oxygen demand can increase rapidly, as illustrated in industrial wastewater and bioremediation studies [12-15].

It was also evident in the affected stretches that there was enrichment in microplastics. Sediment microplastics more than doubled, and biofilm-associated microplastics nearly doubled, illustrating that the canal was a transport channel and retention zone. This is in line with reviews and field studies, which indicate that the distribution of microplastics in freshwater systems is influenced by land use, runoff, sediment deposition, and hydrological setting [16]. The observed absolute sediment load in the present study was comparable with some irrigation ditches, but lower than the reported load in highly polluted rural ditches and urban river sediments [17, 18], which indicates a moderate to high level of local pollution rather than an extreme regional pollution hotspot. The importance of biofilm contamination is that biofilms are living feeding surfaces and early colonization habitats. Experimental and river studies show that biofilms can retain microplastics and transfer them to grazers and protists, turning passive deposits into biologically relevant exposure routes [19, 20]. This, along with data on co-occurring microplastics and metals, helps to establish the ecological relevance of measuring microplastics in both sediments and biofilms [21-23].

Enrichment of both dissolved and fine-sediment fractions in both Zn and Cu suggests multiple exposure pathways for colonizing organisms. Dissolved metals are indicative of direct exposure to the water column, and fine sediments are indicative of an available pool where metals can be bound and are in close contact with the benthic organisms. Also, similar associations have been reported between agricultural or urban drainage and higher concentrations of Cu and Zn in canal-irrigated landscapes [24], drainage sediments [25], and urban-influenced river networks [26]. Furthermore, studies of larger, more dynamic river systems generally find a higher spatial heterogeneity or stronger self-purification than was observed here, pointing to the possibility that the behavior of smaller secondary canals might be that of a depositional sink when flushing and dilution are limited [27, 28]. Lower average concentrations of Cu and Zn, as well as fewer hotspots, have been found in low-disturbance river systems, indicating that the current enrichment is associated with drainage-affected managed canals and not background conditions [29, 30]. Measured concentrations should be considered in the context of published environmental benchmarks and regional background concentrations, not as an indication of ecological risk. A comparison to reported freshwater and sediment reference values indicates that affected reaches had higher concentrations of contaminants, but a full ecological risk assessment would need more data such as contaminant bioavailability, length of exposure, toxicity, and seasonal fluctuations.

Biological response was very strong and consistent between taxonomic and biotic-index metrics. There was a decline in total richness, Shannon diversity, and EPT richness and % EPT, BMWP, and ASPT were the best indicators of ecological impairment. These metrics focus on pollution-sensitive families and sensitive orders of insects [31, 32] and are expected to respond to oxygen stress, fine sedimentation, organic enrichment, and chemical contamination. The decline in sensitive taxa was accompanied by increased representation of tolerant groups, suggesting a shift in community structure rather than a simple reduction in abundance. The observed decrease is consistent with biomonitoring studies conducted in Egypt and other parts of the world, which revealed that polluted canals or rivers exhibited lower BMWP/ASPT and replacement of sensitive taxa with tolerant ones [33, 34]. Depletion of EPT and shifts in tolerance have been observed in urbanizing streams and river systems that are subjected to cumulative stressors [35]. In contrast, hilly streams of high quality or less degraded agricultural ditches may support more diverse EPT or macroinvertebrate communities, demonstrating that canal biodiversity can survive where pollutant stressors and habitat degradation are managed [36]. Although index performance can differ across regions, as seen for non-indigenous biological tools in semi-arid regulated rivers, the direction and magnitude of response in the current study were coherent with the general principles of bioassessment [37].

The functional feeding group findings support the interpretation that the contamination also impacted the functioning of the ecosystem and not just the taxonomic composition. Scrapers and shredders were generally more abundant on reference plates, which are typically indicative of cleaner surfaces, more balanced processing of organic matter, and better oxygen conditions. The collectors dominated the impacted plates, indicative of fine-particle accumulation and suspended organics. This functional shift corresponds to an increase in turbidity, nutrients, microplastics, and sediment-bound metals that are measured. It also further enhances the importance of artificial substrates in engineered canals, where natural-bed sampling might be variable. Hester-Dendy plates standardize surface area and exposure time to help comparisons between reaches in a modified channel [38]. Given the general importance of drains and managed freshwater habitats for biodiversity, not all artificial drains should be excluded from the biological monitoring based on the premise that they are man-made [39].

The integrated stressor-response signal is obtained from the negative regression between the composite pollution index and %EPT. The reduction in %EPT represents a response associated with a combined environmental stress gradient involving microplastics, traces metals, nutrient enrichment, turbidity, and reduced dissolved oxygen. The present design does not allow separation of the independent contribution of individual stressors. This seems ecologically feasible as plastic particles, metals, nutrients, fine sediments, and oxygen decline can operate together in canal habitats. While the field design does not allow separation of the independent toxicity of each stressor, the strong consistency between water chemistry, contaminant burdens, taxonomic metrics, functional groups, and biotic indices provides a strong diagnosis of contaminant-associated impairment.

There are some limitations to note. Although the experimental design provided standardized exposure conditions, the single 28-day deployment represents a snapshot of ecological responses. Temporal changes associated with irrigation schedules, rainfall-driven runoff, and seasonal contaminant transport could influence exposure intensity and should be evaluated through repeated seasonal deployments. Taxa were identified to the lowest practicable level for routine biomonitoring, which may obscure species-level sensitivity. Other stressors, such as pesticides, organic contaminants, and microbial pollution, may also have contributed to the observed response. Nevertheless, the consistency of the verified field, laboratory, and biological data supports the reported contaminant-associated impairment pattern.