pH observations were higher in August and October, at pH 7.78 and 7.77, respectively, while September and November both had pH values of 7.45. In contrast, COD values were lower in August and October at 5.11 and 11.25, respectively. Higher COD observations were observed in September and November, at 29.25 and 25.75, respectively. The high pH observed in August could be attributed to effluent emissions as the flow levels were relatively high. However, as time progressed and temperatures increased, a slight reduction in pH was observed with the onset of the dry season. Furthermore, the flow of the stream was significantly reduced due to some sections of the river being cut off. From September, the site at GRF had completely dried out. Thus, in-stream metabolic processes, such as the formation of fulvic acids rather than effluent emissions could be responsible for the slight decrease in pH observed around September and November. Alternatively, anaerobic conditions resulting from in-stream processes could also be responsible for the variation in COD from September–November.
 

4 Discussion

Different water uses have a different range of variables that define the suitability of water for these purposes. Thus, a ‘holistic’ assessment of WQ will include a range of parameters that, therefore, may limit its use (Bartram & Balance, 1996). For most WQ parameters, benchmarks have been established, which are expected to guide the management of river ecosystems and mitigate the effects of pollution (Enderlein et al., 1988). In this study, WQ was assessed to gain insight into the chemical contamination of Ngwerere River water. Hence, rather than discuss WQ parameters in relation to any specific use of water, the ecosystems approach was applied in assessing the suitability of measuring WQ parameters in surface water. The ecosystems approach is based on objectives for preserving the functional integrity of aquatic systems. Typically, the functional integrity of aquatic systems includes chemical, hydrological, and biological factors and interactions between them (Enderlein et al., 1988).
 
In unpolluted waters, pH is governed by the balance between carbon dioxide, carbonate, and bicarbonate ions and organic acids, such as humic and fulvic acids (de Montety et al., 2011). The natural acid-base balance of a river can be affected by industrial effluent and deposition of acid-forming substances from the atmosphere. Therefore, pH fluctuations can be indicative of river contamination. pH fluctuations over a short duration (24 hours) could be due to photosynthesis and the respiration cycles of algae in eutrophic waters (Menéndez et al., 2001; de Montety et al., 2011). Most natural waters have a pH range of 6.0–8.5. Lower pH values can occur in waters with high organic content, and elevated pH values in eutrophic waters (UNESCO/WHO/UNEP, 1996). The pH values in this study were elevated, within a range of 7.0–8.0, indicating eutrophication. Furthermore, pH observations are consistent with the results obtained by Florescu et al. (2011) for the Arges, Olt, and Jiu Rivers in Romania, which ranged between 7.07–8.5. These ranges are characteristic of surface water systems and are suitable for most organisms (UNESCO/WHO/UNEP, 1996; Florescu et al., 2011).
 
Salinity can profoundly impact riverine ecosystems, resulting in serious environmental issues. Ecosystem degradation due to salinity consequently leads to loss of habitat, biodiversity, native vegetation and water resource value (Nielsen, et al. 2003). Increased salinity reduces the quality of surface water by reducing the levels of dissolved oxygen that life forms depend upon (UNESCO/WHO/UNEP, 1996). According to the Environmental Protection Agency (EPA) (2001), elevated salt concentrations may cause a water system to become unsuitable for domestic, agricultural, or industrial use. Furthermore, according to James et al. (2003), empirical data from ecotoxicological field and laboratory studies attest that there are large variations in salinity thresholds within and between freshwater taxonomic groups. However, generalizations can still be made about salinity thresholds for freshwater biota. The biodiversity of microfauna is inversely related to salinity in freshwater systems (Brock & Shiel, 1983; Halse et al., 1998). Threshold salinity tolerance levels for microinvertebrate species are approximately 2,000 mg l–1(Nielsen et al., 2003). However, according to Hart et al. (1991), Short, et al.(1991), and Kefford (1998), some species can thrive between 5,000 and 10,000 mg l–1. In this study, the salinity ranged between 60–163 mg L-1. These values are below the thresholds for most aquatic biota.
Most authorities such as UNEP, the Water Supply and Sanitation Collaborative Council (WSSCC), or the World Health Organisation (WHO)  have no reference recommendations for WQ thresholds of TDS, (Enderlein et al., 1997). However, attempts to develop TDS standards have been conducted. For instance, the North Dakota State University (NDSU) (2015) proposed TDS recommendations for livestock water. According to NDSU (2015), TDS ≤3,000 mg L-1 is suitable for most livestock, and 3,000–5,000 mg L-1 is suitable for adult livestock, although these levels are detrimental for growing livestock. Levels near 5,000 mg L-1 are intolerable for poultry, while TDS values between 5,000–7,000 mg L-1 are unacceptable for lactating females of most species. High TDS levels of 7,000–10,000 mg L-1 are hazardous to pigs and pregnant or lactating ruminants, including horses, and TDS>10,000 mg L-1 may be fatal. In this study, TDS ranged between 258–567 mg L-1, suggesting that the water is suitable for most livestock species.
 
For many decades, COD has been an expedient variable for assessing chemical pollution arising from different types of waste, such as agricultural (including pesticides and nutrients) and industrial waste (such as heavy metals and persistent organic pollutants), and emerging pollutants (UNEP, 2016). The advantage of COD is that it can be time-saving due to the promptness of measurements (UNESCO/WHO/UNEP, 1996). In unpolluted surface waters, a typical COD range is ≤ 20 mg L-1 O2, while COD will exceed 200 mg L-1 O2 in water that receives effluent. Industrial wastewaters may have COD values that range from 100–60000 mg L-1 O2 (UNESCO/WHO/UNEP, 1996). Given the COD range of 4–36 mg L-1 O2 recorded in this study, and considering the COD limits for unpolluted rivers (UNESCO /WHO/UNEP,1996), COD was indicative of chemical pollution in the Ngwerere River.
 
The elevated Na+ levels in surface water may be caused by various effluents. In coastal zones, the intrusion of sea water can result in elevated Na+ levels in surface water. Thus, sodium levels in surface waters vary according to local geology, sewage, and industrial emissions. Sodium concentrations can range from ≤1–105 mg L-1 or more in natural brines. The WHO set 200 mg L-1 Na+ as the threshold value for sodium in drinking water (UNESCO/WHO/UNEP, 1996). Generally, in surface waters, including those receiving wastewater, sodium levels do not exceed 50 mg L-1 (UNESCO/WHO/UNEP, 1996). High Na+ levels can damage soil structure, resulting in reduced hydraulic conductivity and ultimately causing poor plant growth. The sodium adsorption ratio (SAR) is an estimate of the degree to which Na+ will be adsorbed by the soil and can be used as a measure of the suitability of water for irrigation. High SAR values indicate that the sodium in irrigation water may replace calcium and magnesium ions in the soil, thus damaging soil structure. Increased Na+ will typically result in high SAR values (UNESCO/WHO/UNEP, 1996). The results of this study indicate that Na+ is still below the 50 mg L-1 level expected in many surface waters (UNESCO/WHO/UNEP, 1996).
 
Total suspended solids include materials drifting or floating in the water, which could include silt and sand sediments, plankton, and algae. Decomposing organic matter can also be a constituent of TSS (WHO, 1996; Fondriest Environmental Inc., 2014; Anon, 2017). Many authorities have inexplicit TSS standards for surface water. For example, the best available technique/associated emission levels (BAT/AEL) under Standard 872 of the EU imposes a limit of 5–35 mg L-1 on emissions exceeding 3.5-ton yr-1 (Organisation for Economic Co-operation and Development (OECD), 2007). In Moldova, surface WQ standards stipulate that TSS concentrations at the control point should not exceed the natural levels by more than 0.25 mg L-1 and 0.75 mg L-1 for superior first and second-class fishery water bodies, respectively. In streams with TSS levels above 30 mg L-1 during the low water level period, TSS emissions may be exceeded by 5% for both superior first and second-class fishery water bodies. Discharge of wastewater containing TSS levels of 0.2 mg L-1 for lakes or 0.4 mg L-1 for rivers is prohibited for both classes of fishery water bodies (OECD, 2007). In this study, TSS ranged from 2–710 mg L-1. However, it is a challenge to determine proof of suitability as TSS standards set elsewhere are indirectly stated as rates. Moreover, these rates were set considering the most beneficial ecosystem services in those particular locations, which may be inapplicable to the Ngwerere River.
 

5 Conclusion

This study was undertaken to implement a rapid integrated ecosystem assessment of the Ngwerere River’s predisposition to chemical pollution. To achieve this, the study set out to evaluate selected WQ parameters, determine the spatiotemporal vulnerability of the watershed to chemical pollution, and establish linkages between observed pollution and its potential sources.
 
Findings from this study suggest that there are sources of effluent within the NPW. However, based on the WQ parameters evaluated in this study, the NPW surface water ecosystem is healthy for most ecological functions. This is because some WQ parameters were below the set thresholds, while others were comparatively lower than levels measured in other studies. It was also demonstrated that the position and time of sampling had a significant influence on the variation of some WQ parameters. Salinity and Na+ were significantly influenced by watershed position and sampling site at the p<0.05 and p<0.01 significance levels, respectively. Total dissolved solids and TSS were significantly influenced by sampling site location at the p<0.01 and p<0.001 significance levels, respectively. pH and COD were significantly influenced by the time of sampling at the p<0.01 and p<0.001 significance levels, respectively.
 
Finally, from the findings of this study, it was determined that there was a link between the observed levels of water quality parameters and point and diffuse sources of pollution located in the USC and MSD of the watershed, respectively. To effectively monitor the effects of pollution on freshwater ecosystems, information about existing ecosystem services within river watersheds is required to assist decision-making. Such data is almost non-existent for most watersheds in developing countries, specifically in Zambia. Therefore, there is a need for systematic collection or a repository of data for monitoring and management of freshwater ecosystems. Such repositories will enable the identification of existing ecosystem services, potential pollution threats, and drivers of degradation, as well as physical data for specific watersheds. Further work should be conducted to determine the ecosystem services within the NPW. In this way, ecosystem assessment will be more refined and focus on the environmental concerns of the NPW.