Incorporating Water Quality Analysis into Navigation Assessments as Demonstrated in the Mississippi River Basin

Fig. 2 depicts box plots of monthly mean concentration estimates of NO₃–NO₂, OP, and SSC for the period of record, at the nine sites of interest in the MRB. Sites are ordered from upstream (left) to downstream (right) in the basin (Fig. 1). The center horizontal line in the box plots represent the median values, the boxes show the interquartile range (IQR) between the first and third quartiles (Q1 and Q3), the whiskers extend between Q1 – 1.5 * IQR and Q3 + 1.5 * IQR and the circles indicate values that are potential outliers.

Although there is an established drinking water standard for nitrate of 10 mg/L (USEPA 2021), the desirable level in rivers is less certain. Generally speaking, nitrate levels below about 1.0 mg/L are considered indicative of unpolluted streams in most areas of the world while OP averages about 0.1 mg/L worldwide in unpolluted streams (Meybeck 1982). Fig. 2 shows that sites draining the intensely agricultural regions of the upper MRB (IA-WAPE, IL-VALL, MS-GRAF, and to a lesser degree MS-CLIN) have high nutrient concentrations. The highest levels occur in the small MRB tributaries, such as the Iowa River and the Illinois River. The concentrations in these tributaries lead to elevated concentrations downstream in the mainstream of the Mississippi River at Grafton, Illinois. For NO₃–NO₂, concentrations remain elevated in the mainstem of the Mississippi River even after the addition of the Missouri River, as indicated by higher concentrations at the Mississippi River at Thebes, Illinois, compared to the upstream Missouri River at Hermann, Missouri. NO₃–NO₂ and OP concentrations are lowest in the Ohio River compared to the other sites, and the Ohio River also has comparably higher volumes of streamflow. The two sites farthest downstream in the MRB (MS-STFR and AC-MELV) have NO₃–NO₂ and OP concentrations that are slightly higher than those in the Ohio River, with means around 2 and 0.1 mg/L, respectively. Approximately 23% of the flow of the lower Mississippi River is diverted into the Atchafalaya River (Allison et al. 2012), accounting for some of the similarity in NO₃–NO₂ and OP concentrations between these two sites (Fig. 1).

Seasonally, NO₃–NO₂ concentrations tend to be highest in the spring and early summer, perhaps as snow melt, spring rains, and fertilizer applications combine to cause excess in runoff from fields in agricultural regions, and then concentrations rapidly decline during the drier mid-summer and fall (Fig. 2). For many of the sites, NO₃–NO₂ concentrations remain elevated during winter, especially in the upper MRB. Seasonal patterns for OP are less consistent between sites. OP concentration at mainstem sites on the Mississippi River tend to follow a sinusoidal shape with highest concentration occurring mid-summer through early fall and lowest concentrations during the winter. OP in the Iowa, Illinois, and Ohio Rivers tend to have similar pattern though shifted to later in the year with high concentrations during the winter and lower concentrations in spring and summer. OP in the Missouri River is similar to the mainstem Mississippi River sites but concentrations are fairly uniform throughout the year with a slight increase during the summer.

Suspended sediment concentrations have a very different spatial pattern than nutrients do across these nine sites in the MRB. The upper Mississippi River and three tributary sites have relatively low SSC concentrations with a few outliers, likely due to large storms or floods. Conversely, the Missouri River has very high SSC concentrations with many high outliers. The contributions of the Missouri River strongly influence SSC concentration in the mainstem Mississippi River at Thebes, Illinois, but below the relatively sediment-free Ohio River, concentrations are much lower. A slight increase in SSC is seen in the Atchafalaya River, after its diversion through the Old River Control structures, suggesting a within-basin source of suspended sediment for the Atchafalaya River that is absent in the lower Mississippi River, or contributions from the Red River, which may have implications for the restoration of deltaic wetlands in coastal Louisiana (Mize et al. 2018).

In Fig. 3, box plots of monthly mean flux estimates of NO₃–NO₂, OP, and SSC, as a rate of metric tons per day, are shown for the nine sites in the MRB. Due to the often-overwhelming influence of streamflow magnitude on flux estimates, the flux patterns in Fig. 3 are markedly different from the concentration patterns in Fig. 2, due largely to differences in flow volumes between the sites. The smaller tributary rivers, such as the Illinois and Iowa Rivers, which have very high nutrient concentrations, have relatively small streamflows such that their fluxes are a small portion of the overall nutrient load of the Mississippi River. However, in aggregate the upper MRB, as captured at the Mississippi River site at Grafton, Illinois, has a large nutrient flux compared to the large downstream tributaries, the Missouri and Ohio Rivers. This pattern of nutrient accumulation to the total flux continues moving downstream on the Mississippi River with additional inputs from the Missouri and the Ohio Rivers, as well as other tributaries which have less robust historical measurements between the Ohio River and lowermost study site (MS-STFR), such as the Obion River, the White River, and the Yazoo River. Despite its very large drainage basin size and high nutrient concentrations, the flow of the Missouri River is small enough that the fluxes of nutrients from the Missouri River are relatively low. Conversely, the modest nutrient concentrations in the Ohio River become a relatively large nutrient flux with the river’s large flow volumes. The seasonality of nutrient fluxes is controlled by the hydrologic seasonality of the system and likely by the application of agricultural fertilizers, both peaking in the spring and early summer.

Considering flux of SSC, it is easy to understand the old adage “in the lower Mississippi River, the sediment comes from the Missouri River, but the water comes from the Ohio River.” The SSC flux from the Missouri River dominates the seasonal pattern seen at the Mississippi River at Thebes, Illinois (Fig. 3). Much less suspended sediment enters the system from the Ohio River, resulting in considerably lower fluxes at St. Francisville, Louisiana, compared to those upstream in the Missouri River or the Mississippi River at Thebes. Upstream from St. Francisville, Louisiana, about 23% of the SSC flux in the Mississippi River is diverted into the Atchafalaya distributary (Allison et al. 2012), which also contributes to some of the decline in SSC flux in the Mississippi River between Thebes, Illinois (MS-THEB) and St. Francisville, Louisiana (MS-STFR). However, given the magnitude of sediment fluxes in the Mississippi River at Thebes, Illinois, the contributions of sizable tributaries downstream (such as the Arkansas, Hatchie, and Yazoo Rivers, which are not well measured), and the influx of the Ohio River, the diversion of some of this sediment load into the Atchafalaya distributary likely does not fully account for the lower sediment flux in the Mississippi River at St. Francisville, Louisiana. This apparent loss of SSC flux could be influenced by a variety of factors, including the possible deposition of these fine sediments in the floodplain and other sinks along the lower Mississippi River (Allison et al. 2012; Little and Biedenharn 2014). One shortcoming this observation highlights is the lack of sites on the Mississippi River with long-term nutrient and SSC records between Thebes, Illinois, and St. Francisville, Louisiana (Fig. 1). A multiagency effort to establish new data collection sites and to better utilize existing data could improve the spatial coverage of data and help refine the nutrient and sediment budgets for this section of the river. If better quantified, the aggregational tendency of sediments in channel and floodplains in the lower Mississippi River might be more evident and could show potential effects on navigation, flood risk management, and coastal restoration in the future.

An additional method to understand the water quality at a site is to evaluate changes over time in concentration and flux with some of the variability related to fluctuating weather and streamflow removed. In Figs. 4–6, modeled trends in monthly mean flow-normalized concentration and flux are shown for NO₃–NO₂, OP, and SSC for the Iowa River, Missouri River, and the lowermost site on the Mississippi River (MS-STFR). The trends, which smooth out some of the influence of variable year-to-year streamflow volumes on river water quality, are graphed for four months (February, May, August, and November), representing each of the four seasons (winter, spring, summer, and fall, respectively), to examine seasonal changes over time.

Fig. 4 features the Iowa River at Wapello, Iowa, and shows high concentrations and a distinct peak in NO₃–NO₂ concentrations and fluxes in all seasons, in the early 2000s, followed by a gradual decline. The winter and spring show relatively stable conditions during the last part of the record, whereas the summer and fall show slight increases during this time. Since 1980, long-term changes in NO₃–NO₂ concentrations were distinct among seasons. A large decrease occurred during the fall, smaller decrease during the winter, a large increase during the spring, and fluctuating but little change during the summer. Changes in OP concentrations [Fig. 4(a)] over time show decreases until the mid-1990s, followed by increases in all seasons with a distinct peak in OP concentration in the mid-1990s during the winter. Early 2000s show a peak in OP flux, especially during the winter and spring [Fig. 4(b)]. SSC concentrations and flux peaked in the 2010s, especially during summer months, but have been gradually declining since that time (Fig. 4).

In contrast, concentrations are nearly four times less and trends are more uniform across seasons for both NO₃–NO₂ and OP concentrations and flux on the Missouri River at Hermann, Missouri; however, both have been steadily increasing over time (Fig. 5). Concurrently, SSC has been basically stable during the low flow winter months since the mid-1980s and significantly declining during the spring and summer months (Fig. 5).

Following a rise in NO₃–NO₂ concentrations and fluxes in the early 1980s (Fig. 6), the mainstem of the Mississippi River at St. Francisville, Louisiana, has been fairly stable with regard to NO₃–NO₂ during the winter and spring, while the summer and fall show some increases, particularly during the last 10–20 years. OP concentrations and fluxes had similar trends across seasons with declines during the 1980s, followed by increases over the following years (Fig. 6). Conversely, SSC concentrations and fluxes decreased dramatically during the several decades, particularly in the winter and spring, following the sediment trends reported by (Mize et al. 2018) for this location. Due to the extremely large flow volumes in the lower part of the Mississippi River and related dilution effects, this stability over time in the lower river for NO₃–NO₂ is not surprising, but even within this context, there is a clear indication of increases in OP but declines in SSC concentrations and fluxes (Fig. 6).

While Figs. 4–6 allow a visual comparison of seasonal trends across the three constituents at three of the sites, Fig. 7 shows a comparison of annual trends across all nine sites. The trend estimates are reported as the annualized percent change in flow normalized concentration or flux between the start and end years of three trend periods. These trend estimates are binned (i.e., color coded on Fig. 7) using the p value from a bootstrap test for trend direction. Trend estimates with p values <0.05 were considered upwards or downwards depending on the trend direction. Trend estimates with p values ≥0.05 but <0.20 were considered to have some evidence of upward or downward trends, and trend estimates with p values >0.20 were considered to have uncertain or little change. This analysis covers roughly a 40-year period from WYs 1980 or 1982 to WY 2019. Two subperiods were also analyzed: a 30-year period (or less) from WY 1980 (or WYs 1982 or 1997, depending on the constituent and site) through WY 2010, which may be compared to the nitrate analysis in Murphy et al. (2013), and a period focused on the most recent decade from WY 2010 to 2019. For NO₃–NO₂, the earliest trends start in WY 1980, for OP they start in WY 1982, and for SSC it depends on the site with some not starting until WY 1997 (Table 1). Because of the varying length of the trend periods by constituent and site, Fig. 7 reports the annualized percent change.

Starting at the northern most site in this analysis, the Mississippi River at Clinton, Iowa, had NO₃–NO₂ concentrations that were increasing at about 1.8 percent per year (%/year) over the period of record, but during the last decade changes in concentration began to level off (Fig. 7). Fig. 4(a) showed that the seasonal NO₃–NO₂ concentration trends varied at the Iowa River, with strong increases during the spring driving the overall increase annually over the period of record. However, there has been little change over the last decade (Fig. 7). NO₃–NO₂ concentrations at the Illinois River at Valley City, Illinois, have decreased throughout the period with less change happening during the most recent 10 years. The Mississippi River at Grafton, Illinois, which is the first site downstream from the confluence of the Illinois and Mississippi Rivers, exhibits relatively stable conditions over time, integrating conditions from the upper Mississippi River and the Illinois River. As with Fig. 5, the Missouri River at Hermann, Missouri, exhibits markedly different temporal patterns from the other sites showing increasing NO₃–NO₂ concentrations, at a rate of 2.5%/year throughout the study period, with particularly strong increases in the last decade (4.3%/year). The volume of the Mississippi River moderates the influence of the Missouri River, so NO₃–NO₂ concentrations at the Mississippi River at Thebes, Illinois, have remained fairly stable. NO₃–NO₂ concentrations have gradually declined at the mouth of the Ohio River, though the direction of change is uncertain, and concentrations are stable to decreasing in the lower Mississippi River and the Atchafalaya River, with some uncertainty.

OP concentrations and flux in the MRB are largely increasing at most of the nine sites (Fig. 7). While total phosphorus is commonly associated with fine sediments, OP, which is defined as the dissolved, bioavailable component of phosphorus, might be expected to be more similar to NO₃–NO₂. However, while NO₃–NO₂ trends varied depending on the site, OP increased at many of the sites for the first 30 years of record. OP continued to increase at some of these sites during the most recent 10 years, and at other sites, new increases occurred during this time that were not evident previously in the record. The exception is the Illinois River, where after a long period of increases in OP concentration, the last decade has shown some evidence of a decrease, and this trend is reflected downstream in the Mississippi River at Grafton, Illinois, though with less certainty. In the case of the Illinois River, NO₃–NO₂ and OP concentrations changed little or decreased, respectively, from WY 2010 to 2019, suggesting efforts to moderate nutrient inputs to the river may have been successful. Conversely, the Missouri River at Hermann, Missouri, showed large likely increases in NO₃–NO₂ and OP concentrations and flux. While not exhibiting significant NO₃–NO₂ trends for most of the trend periods, the Ohio River and the two sites in the lower MRB, show increasing OP concentrations and flux over the same period, but with much of the change occurring during the last decade.

Concentration and flux trends for SSC for the time periods available show widespread decreases across the sites (Fig. 7). Almost all nine sites show decreases in SSC concentration and flux from WYs 1980 or 1997 to WY 2019, typically at a rate of around −1.5%/year. Changes in SSC may have leveled off or slowed in the last 10 years in the Illinois River, with an annualized change of about −0.5%/year, and the Missouri River, with an annualized change of about 0.04%/year. Interestingly, these downward trends in SSC are evident at the lowermost Mississippi River site at St. Francisville but are not as strong in the Atchafalaya River at Melville, Louisiana.

Figs. 8 and 9 are an illustration of a useful water quality analysis graphic generated using EGRET, the R package used to run WRTDS. These contour plots illustrate the expected (modeled) concentration of a constituent over time, in relation to the range of streamflow volumes that occurred at that site over the period of record. The black lines on the plots indicate the 5th and 95th percentiles of streamflow and their periodic oscillation shows how these magnitudes of streamflow vary over the course of a year with lower flows in the winter, increases in flow in the spring and summer, and decreases in the fall. From Fig. 8, we can ascertain that the highest NO₃–NO₂ concentrations in the Iowa River during the last 40 years occurred during the 1980s at moderate to high flow conditions during the winter. Concentrations in the Iowa River are still high but have decreased particularly at moderate and high flows, as illustrated by the gradual lightening of color across the x-axis. High concentrations primarily occur during high flows. Conversely, the highest NO₃–NO₂ concentrations on the Missouri River at Hermann, Missouri (Fig. 9), are nearly four times lower than the Iowa River, and the highest concentrations have occurred during the most recent five-year period at moderate and low flows. This plot supports the trends from Fig. 6 showing increases in NO₃–NO₂ concentrations over time and indicates that some relatively high concentrations occur during low to moderate flows, particularly during the winter and spring. A comparison of these two plots suggests that the source of nitrate for these two rivers is very different and that different solutions for improving nutrient levels are likely necessary.