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a Annis Water Resources Inst., Grand Valley State Univ., 740 West Shoreline Dr., Muskegon, MI 49441
b Soil and Water Science Dep., Inst. of Food and Agric. Sci., 106 Newell Hall, P.O. Box 110510, Gainesville, FL 32611-0510
* Corresponding author (steinmaa{at}gvsu.edu)
Received for publication February 23, 2004.
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ABSTRACT |
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 TOP ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
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Abbreviations: DO, dissolved oxygen • SRP, soluble reactive phosphorus • TP, total phosphorus
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INTRODUCTION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
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Mineral associations play an important role in the release of P during anoxic or anaerobic conditions. Phosphorus associated with iron minerals can become soluble in the absence of oxygen while the P fraction associated with calcium-based minerals may remain stable (Mortimer, 1971). Phosphorus release rates have been found to be closely correlated to the iron-bound fraction in sediment (Petticrew and Arocena, 2001).
Historically, field measurements of P concentration in lakes have concentrated on the "external" loading of P. This is the contribution of P from point and nonpoint sources flowing into a waterbody. The contribution of P being released from the sediments, or the internal load, while acknowledged to be of potential importance, is measured less often. However, in eutrophic lakes, internal loading can account for a substantial amount of the total P load (Moore et al., 1998). Indeed, many studies have shown that reductions in external loading, to levels where water quality improvement should be detected, do not have the desired effect because of the counteracting release of P from sediments (Björk, 1985; Graneli, 1999; Steinman et al., 1999).
Although many sediment management technologies exist to deal with internal loading, the most common practices include chemical treatment, oxidation, and dredging (Cooke et al., 1993). Chemical applications are intended to bind the P, and usually include aluminum sulfate (alum), lime, or iron (Cooke et al., 1993). Alum is particularly effective due to its dual mode of action for P removal. Alum reacts with soluble P to form an insoluble precipitate (Stumm and Morgan, 1996). In addition, alum will form an insoluble aluminum hydroxide floc at pH 6 to 8, which has a high capacity to adsorb large amounts of inorganic P (Kennedy and Cooke, 1982). By these two mechanisms, an alum application can irreversibly bind P and inhibit diffusive flux from sediments.
Previous studies showed that the effectiveness of alum application depends on a number of factors. For example, Welch and Cooke (1999) reported that alum application reduced internal loading in seven of seven dimictic lakes by an average of 80% during a period of 4 to 21 yr (avg. length of effectiveness: 13 yr). However, they also found that internal loading was reduced in only six of nine polymictic lakes, with reductions averaging 67% and remaining effective for an average of 10 yr. They speculated that differences in macrophyte density (which can physically interfere with floc distribution), rates of plant senescence, and/or resuspension of sediments by bioturbation may have been responsible for the different results both among and within lake types.
Studies on hypereutrophic lakes in Florida indicated that internal loading could contribute a significant amount of bioavailable P in these lakes (Moore et al., 1998; Steinman et al., 1999). The source of P in these cases was the top layer of sediment. If P is not tightly bound to sediment, it becomes available to the water column on resuspension of the surface sediments or simply by gradient flux under the appropriate conditions of temperature, pH, dissolved oxygen, and ambient surface water P concentration. If there is a similar significant internal loading source of P in Spring Lake, reductions in external P loads alone likely will be insufficient to reduce P levels in the lake water, at least for the foreseeable future. Although there is general consensus that internal load will eventually decline following external load reduction, the time required may be long and rates may actually increase temporarily (Ahlgren, 1988; Sas, 1989; Chapra and Canale, 1991; Søndergaard et al., 1993; Welch and Cooke, 1995).
Spring Lake faces some of the most critical water quality challenges in west Michigan. The TP concentrations in Spring Lake are usually far in excess of water quality standards. For example, the USEPA has set a surface TP water quality goal of 0.015 mg L–1 for the west Michigan ecoregion (USEPA, 2000). However, during ice-free periods from 1999 through 2002, surface TP concentrations in Spring Lake averaged 0.100 mg L–1 (range: 0.006–0.631 mg L–1).
The goals of this project were to measure whether internal loading of P from the sediment to surface water is a significant source of P to Spring Lake, and if alum application to the sediment surface was effective at controlling P release.
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MATERIALS AND METHODS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
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Sediment cores were collected using a piston corer (Fisher et al., 1992). Twelve cores were collected from each site (24 per sampling date). The coring device was constructed with a graduated 0.6 m long polycarbonate core tube (7 cm i.d.), aluminum drive rods, and a PVC attachment assembly for coupling. The piston was advanced 20 cm before deployment to maintain a water layer on top of the core during collection. The corer was vertically positioned at the sediment–water interface and pushed downward with the piston cable remaining stationary. After collection, the core was brought to the surface and sealed with a rubber stopper before removal from the water, resulting in intact sediment cores that were approximately 20 cm in length, with a 25-cm overlying water column. The piston was then bolted to the top of the core tube to keep it stationary during transit. Core tubes were placed in a vertical rack and transported to the laboratory in a water bath to keep temperatures near ambient.
Laboratory Set-Up and Analysis
The 24 sediment cores (12 per site) collected on each sampling trip were placed into a darkened Revco environmental growth chamber (Revco Scientific, Asheville, NC), with the temperature maintained to match ambient conditions in the field (see Table 1). The water column of the sediment cores from each site was exposed to one of four treatments (three replicates per treatment per site): (i) aerobic with alum, (ii) aerobic without alum, (iii) anaerobic with alum, and (iv) anaerobic without alum. Nitrogen (with 330 mg L–1 CO2) or air was bubbled into the water column of each tube to create aerobic or anaerobic conditions, respectively. Aluminum sulfate solution [Al2(SO4)3·14H2O] was added to half of the aerobic and half of the anaerobic water column treatments at a concentration of approximately 25 mg Al L–1 (
6.6 g Al m–2). At this concentration, a 1-cm deep floc formed at the water–sediment surface. The dosing concentration was based on the recommendations outlined in Cooke et al. (1993). Alum application was considered time zero. Alum was obtained from General Chemical Corporation (River Rouge, MI).
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Flux calculations were based on the increase in water column concentrations of TP. Calculations were based on three different time periods to reflect: (i) the maximum release rates (linear portion of the curve), (ii) moderate release rates, and (iii) minimum release rates. These different rates allowed us to capture the full range of potential internal loading rates, and gain a better understanding of the possible uncertainties in estimating internal loading in Spring Lake. Phosphorus flux was calculated using the following equation:
 | [1] |
Internal load at a specific site was calculated by scaling up the mean P flux from the ambient (i.e., anaerobic/no alum) treatment condition as calculated in Eq. [1] to the entire lake area and multiplying by the percentage of time during the year that the lake was estimated to have DO concentrations <1 mg L–1 (see below).
Following the incubations, cores were centrifuged to remove excess porewater and the top 10 cm of each core was sequentially fractionated (Moore and Reddy, 1994) to determine the fraction of P bound to Fe/Al and Ca/Mg in the sediments. Residual sediment was shaken for 17 h with 0.1 M NaOH and centrifuged, and the porewater was then filtered and analyzed for SRP. This fraction represents the Al- and Fe-bound P. After this extraction, the residual sediment was extracted for 24 h with 0.5 M HCl and centrifuged, and the porewater was filtered and analyzed for SRP. This fraction represents the Ca- and Mg-bound P.
Phosphorus analyses were performed on a BRAN+LUEBBE Autoanalyzer (Bran+Luebbe, Delavan, WI) by the automated ascorbic acid method (USEPA, 1983). Sediment extracts were neutralized before analysis.
Statistical Analysis
Phosphorus release rates were calculated for individual cores and treatments compared (n = 3) within each site using analysis of variance. Tukey's post-hoc multiple comparison test was used to determine if mean release rates from individual treatments were significantly different from one another. Phosphorus concentrations of the chemically fractionated sediment cores also were analyzed by analysis of variance to detect statistically significant differences among sites and treatment conditions. All statistical analyses were conducted using SAS (version 8).
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RESULTS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
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To determine the frequency of anoxia in Spring Lake, historic dissolved oxygen data in Spring Lake, collected as part of Grand Valley State University's vessel education program, were analyzed. Cruises onboard the D.J. Angus from 1998 through 2002 resulted in a total of 864 sampling events. These data revealed that DO was <2 mg L–1 between 10 and 31% of the time and <1 mg L–1 between 4 and 25% of the time (Table 2). These data provide a general idea of anoxic frequency in Spring Lake, but they must be treated with caution. First, they were collected from only one, relatively deep station. As a consequence, they likely overestimate the percent of time low-DO concentrations exist in the lake, especially at shallower stations. Second, they are snapshots, taken only during daytime cruises. If Spring Lake exhibits diurnal cycles in anoxia, with low-DO conditions more likely at night, these data will not capture this phenomenon and therefore underestimate anoxic conditions. Finally, data were collected only from late April through early October. We assume aerobic conditions during the remainder of the year due to greater mixing of the water column and reduced metabolism in the benthos because of colder temperatures, but it is also likely that anoxia develops during periods of ice cover. If anoxic or hypoxic conditions do develop during these months, the amount of low-DO conditions in Spring Lake would be underestimated.
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10% (Site 4) to 10-fold (Site 2; Table 4). For the anaerobic/no alum treatment, "low" TP release rates ranged from about 1.6 (Site 2) to 14.8 (Site 1) mg P m–2 d–1, whereas "high" TP release rates ranged from about 12 (Site 4) to 29.5 (Site 1) mg P m–2 d–1 (Table 4). The "high" TP release rates were greater at Sites 1 and 2 than at Sites 3 and 4 (Table 4).
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Internal Load Calculations
Internal TP loads were estimated from the anaerobic/no alum treatment. This treatment reflected the condition when P release was most likely to occur in nature. Internal P loads varied approximately 2.5-fold from the mean in the low range to the mean in the high range (Table 5). Hence, the portion of the curve used to estimate P release rates clearly has a significant impact on calculating the internal load in Spring Lake. Overall, the estimated internal P load to Spring Lake ranged from a low of 0.6 Mg yr–1 at Site 2 to a high of 10.6 Mg yr–1 at Site 1. The mean internal P loads were 2.7, 6.2, and 6.4 Mg yr–1 under the low, medium, and high categories, respectively (Table 5).
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Sediment Phosphorus Forms
Mean NaOH-extractable inorganic P (NaOH–Pi) concentrations (i.e., Fe/Al-bound P) from the sediment cores ranged from approximately 123 to 200 mg kg–1 dry weight (Fig. 4)
. There were no statistically significant differences among sampling location, redox state (aerobic vs. anaerobic), or alum treatment (present vs. absent) for the NaOH–Pi fraction. Mean HCl-extractable P (HCl–Pi) concentrations (i.e., Ca/Mg-bound P) were more variable than NaOH–Pi, ranging from 126 to 513 mg kg–1 dry weight (Fig. 4). With the exception of Station 3, the HCl–Pi concentrations were greater than NaOH–Pi concentrations. A statistically significant difference between sites was noted (p = 0.01) for HCl–Pi, with Stations 3 and 4 being significantly lower than Stations 1 and 2. As with the NaOH–Pi, no statistically significant differences were observed with respect to redox or alum treatments for the HCl–Pi.
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DISCUSSION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
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The P release rates from Spring Lake were in the same range as those measured in eutrophic systems (6–18 mg m–2 d–1; Nürnberg and LaZerte, 2004), and even approached some of the highest recorded release rates (30–60 mg m–2 d–1; Nürnberg, 1988). These results strongly suggest that internal P loading is a significant source of P in Spring Lake, potentially accounting for somewhere between 55 to 65% of the total P load to the system. However, a number of assumptions were built into these calculations, which are evaluated below.
First, it was assumed that release rates from sediments in the core tubes were representative of sediments and conditions in Spring Lake. The sampling strategy was designed to cover as much of the geographic range in Spring Lake as possible. Indeed, this study revealed that there was spatial variability in P release rates in Spring Lake; in general, P release rates were higher at Sites 1 and 2 than at Sites 3 and 4. It does not appear that this difference was due to sediment chemistry, as the Fe-bound P sediment fractions were similar among all sites. Although the Ca-bound P sediment fraction was higher at Sites 1 and 2, if anything this should account for lower, not higher, P release rates. Since our sampling approach introduced no systematic bias, we have no reason to believe our results are not spatially representative of Spring Lake as a whole.
The second major assumption in our calculations was that the measured P release rates applied whenever DO concentrations were <1 mg L–1. Under aerobic conditions, oxidized iron will remain bound to P in the sediments but when conditions become anaerobic, the reduced form of iron becomes more soluble, and the P is released (Mortimer, 1941, 1942). In addition, biological processes (e.g., bacterial activity, mineralization processes, and bioturbation), chemical parameters (e.g., pH, alkalinity, and nitrate), and physical factors (e.g., resuspension and sediment mixing) also will influence P release rates (Boström et al., 1982; Søndergaard et al., 1992; Petterson, 1998). We used a 1 mg L–1 threshold for DO (cf. Mortimer, 1971; Nürnberg, 1995); this was an operational threshold, as our datasonde was actually located approximately 1.0 m above the sediment–water interface, and it is likely that conditions were even more reduced below this depth.
The third assumption dealt with the calculation for percentage of year that DO concentrations in Spring Lake were <1 mg L–1. This extrapolation was based on data from one relatively deep site ("Deep Hole") in Spring Lake, near Site 1. It is likely that the DO values at "Deep Hole" were lower, on average, than at other sites in Spring Lake due to its greater depth, which would result in an overestimate of anaerobic conditions (and internal loading). However, this overestimate is offset, at least to some degree, by the assumption that only oxic conditions existed between October and April (when no DO samples were taken). In fact, it is likely that anoxic conditions do occur under ice cover, at least occasionally (thereby releasing P). However, given the absence of data, a conservative approach was adopted by assuming continuous oxic conditions in the winter.
Fourth, calculations of P release rates under the moderate and especially the low ranges should be viewed with caution. Although inclusion of these rates are of value because they provide estimates of the lower and middle ranges of P flux, the inclusion of data from longer periods of incubation increase the risk that chemical or biological interference may affect the rates.
Finally, it was assumed that the incubation conditions were representative of natural conditions. Although the laboratory conditions mimicked the ambient temperature and light regime, clearly the hydrodynamics were altered. It is likely that the laboratory set-up for the anaerobic water column with no alum treatment represented an optimal situation for release of P (constant anaerobic conditions) compared with natural conditions. In addition, the P concentration of our replacement water (filtered lake water) was lower than the P concentration of the water remaining in the core tube, especially in the anaerobic/no alum treatment (which accumulated P over time). This introduction of relatively low P concentration water into the water column steepened the concentration gradient inside the core tubes, likely enhancing P release. As a consequence, these release rates should be viewed as maximum potential rates.
Given the spatial variability in P release rates among sites, it was surprising that the extractable P in the sediments was poorly related to site, redox state, or alum treatment. Previous studies have reported a reduction in the iron-bound P fraction due to the reaction with the alum layer (Kennedy and Cooke, 1982; Cooke et al., 1993). It is possible that our 10-cm deep sediment cores biased the results by introducing sufficient additional P to mask chemical changes from the alum treatment or the differences in redox conditions.
Management strategies to control internal loading usually include sediment removal and chemical applications, such as Al, Fe, or Ca salts (Cooke et al., 1993). The data from this study clearly showed that alum application was very effective at reducing internal P loading rates in our sediment cores. Irrespective of location or oxic state of the treatment, TP release rates were virtually negligible when alum was applied. Alum is particularly effective due to its dual mode of action for P removal. When added to water at pH 6 to 8, alum dissociates to give trivalent Al3+ ions, which undergo a series of rapid hydrolytic reactions to form soluble monomeric and polymeric species, as well as an amorphous Al(OH)3 floc (Bottero et al., 1980; Omoike and Valoon, 1999). The monomeric species are capable of precipitating soluble P as Al(PO4). In addition, the amorphous Al(OH)3 floc can remove soluble and particulate forms of P by adsorption and/or physical entrapment (Galarneau and Gehr, 1997). The alum floc will settle and form a layer at the sediment–water interface that has a high capacity to adsorb large amounts of inorganic P (Kennedy and Cooke, 1982).
While an alum treatment is likely to have short-term benefits, it is unclear how long an alum treatment would be effective in Spring Lake. The current study was not designed to address the question of long-term effectiveness of an alum treatment. However, prior studies have shown that effectiveness ranges from 4 to 20 yr, and is dependent on many factors, including: (i) the morphometry of the lake, which influences the likelihood that the alum will be resuspended by wind–wave action, and no longer covers the sediments uniformly (Welch and Cooke, 1995, 1999); (ii) the amount of alum added to the sediment, to ensure there is sufficient aluminum to bind the P, but not add more than necessary because of financial or environmental concerns (Rydin and Welch, 1998; Lewandowski et al., 2003); (iii) activity from bottom-swelling animals (i.e., bioturbation) in the sediments, which can enhance P flux from the sediments due to particle mixing and alteration of the redox conditions (Van Rees et al., 1996; Matisoff and Wang, 1998). In addition, bioturbation can redistribute and bury the alum, reducing its efficacy; (iv) presence of macrophytes, either by intercepting the alum floc and preventing a uniform cover over the sediment or by P release from tissue during plant senescence (Welch and Schrieve, 1994; Welch and Cooke, 1999); (v) water column pH, as circumneutral waters (pH 6–8) are optimal for creating an alum floc (Rydin and Welch, 1998, Lewandowski et al., 2003); (vi) rate of sedimentation in the water column because new organic matter that settles over the alum can reduce its ability to bind P (Lewandowski et al., 2003); and (vii) the influence of internal loading from shallow areas not treated by alum. Significant internal loading has been reported in shallow lakes and areas where frequent mixing occurs (Nixdorf and Deneke, 1995; Søndergaard et al., 1999).
A vital prerequisite for restoring lake water quality is the removal of the underlying reasons for the impairment. Thus, regardless of the long-term effectiveness of an alum treatment, it is critical that external load reduction complement any chemical addition (Hansson et al., 1998). Continued efforts at reducing stormwater discharge, conversion of septic systems to sewers, use of low-P fertilizer, and implementation of other best management practices should be emphasized, along with the provision of appropriate incentives in the Spring Lake watershed.
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ACKNOWLEDGMENTS |
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REFERENCES |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
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