Natural Resources Research Institute, University of Minnesota-Duluth, ... 1 Present address: Department of Environmental and Technological Studies, St. Cloud ...
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Hydrobiologia 361: 145–156, 1998. c 1998 Kluwer Academic Publishers. Printed in Belgium.
Uses of alkaline phosphatase activity in evaluating phytoplankton community phosphorus deficiency Charles Rose1 & Richard P. Axler Natural Resources Research Institute, University of Minnesota-Duluth, Duluth, MN 55811, U.S.A. 1 Present address: Department of Environmental and Technological Studies, St. Cloud State University, St. Cloud, MN 56301, U.S.A. Received 16 September 1997; in revised form 6 November 1997; accepted 6 November 1997
Key words: Alkaline phosphatase activity, growth bioassays, Lake Superior, phosphorus deficiency, phytoplankton
Abstract The phosphorus (P) deficiency status of phytoplankton communities was measured using the physiological indicator, alkaline phosphatase activity (APA) and nutrient-addition growth bioassays in field sampled from four northeastern Minnesota lakes and the far western arm of Lake Superior. Phosphorus additions generally reduced APA, while other treatments increased activity. Samples receiving nitrogen (N) and P increased APA after a long lag period. P-addition bioassays of Lake Superior were consistent with phytoplankton P limitation and variations in APA indicated potential seasonal and spatial changes in P deficiency status. The results suggest that APA reliably reflected the phytoplankton P status, but may not provide sufficient information when N or NP limitation is present. Introduction The phosphatase group of enzymes hydrolyze ester bonds between phosphates and dissolved organic molecules, making phosphates available for cellular assimilation. Phytoplankton-associated alkaline phosphatase activity (APA) has been used as an indicator of phosphorus (P) deficiency because it is synthesized at low levels of P availability (Pettersson, 1980), repressed when P becomes available (Perry, 1972; Elser & Kimmell, 1986) and has been shown to be inversely related to extracellular and intracellular P concentrations (see Jansson et al., 1988 for summary). However, it is difficult to distinguish among phosphatases produced by other organisms such as zooplankton (Jansson, 1976; Boavida & Heath, 1984) or bacterioplankton (Cembella et al., 1984; Jansson et al., 1988), and there may also be a background level of constitutive enzymes. Another limitation of its use as an indicator of P-deficiency is that substrate availability is likely to limit the reaction’s rate in nature, not the enzyme concentration. Despite these concerns, APA has been shown to be related to algal P limitation in natural systems (Fitzgerald & Nelson, 1966; Berman, 1970; Healey & Hendzel,
1979, 1980; Smith & Kalff, 1981; Vincent, 1981; Gage & Gorham, 1985; St. Amand et al., 1989; Istvanovics et al., 1992; Vrba et al., 1995 and others). Enzymatic assays may also avoid container and incubation artifacts associated with nutrient enrichment bioassays (Elser & Kimmel, 1986). Jansson et al. (1988) suggested that APA may be used as a P-deficiency indicator when the origins of the enzymes are addressed and when used in conjunction with other indicators. Because most APA methods use artificial substrates under optimized conditions, and the in situ pool of hydrolyzable organic phosphorus is difficult to determine (Jansson et al., 1988), they are generally used as a P-deficiency indicator or to assess spatial and temporal change rather than as an in situ estimate of inorganic phosphate regeneration. This study was designed to determine if APA patterns were consistent with expected responses during changing nutrient status (due to nutrient additions) in phytoplankton growth bioassays, and how APA can be used to indicate changes in phytoplankton nutrient status in a lake with strong P limitation (Lake Superior).
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146 Methods Field sampling
Figure 1. Sampling sites in northeastern Minnesota and the far western arm of Lake Superior.
Study area Four northeastern Minnesota lakes were used to provide lakes with differing trophic status and nutrient limitation and to take advantage of existing state monitoring programs (see Figure 1). Thrush Lake is a 6.62 ha oligotrophic, headwater lake. Dunnigan is a 34.6 ha, mesotrophic lake and Divide is a 23.4 ha, mesotrophic headwater lake. The three lakes are located in the Superior National Forest. The fourth lake, Sand Lake, is a 49.8 ha, shallow, meso-eutrophic, seepage lake surrounded by agricultural land. These lakes were part of the Minnesota Pollution Control Agency’s Acid Deposition Monitoring Program. Further information regarding these lakes is contained in Axler et al. (1991a and 1994). Figure 1 also shows that the primary Lake Superior sampling site (LR3) used for APA experiments was located about 5 km east of the mouth of the Lester River (46 48.30’ N, 91 58.20 W). Other sampling stations were located along a transect from the mouth of the Lester River to near the center of the far western arm of Lake Superior (LR1, LR2, and LR5), near the Gooseberry River mouth (GR at 47 090 N, 91 270 W), and near the Knife River (KR at 46 56.40 N, 91 31.90 W). Addition sites in the far western arm of Lake Superior were also sampled. While phytoplankton productivity in Lake Superior has consistently been shown to be P-limited (Schelske et al., 1972; Plumb & Lee, 1974; Shapiro & Glass, 1975; Fahnenstiel et al., 1990), studies of the other four lakes have shown that their phytoplankton may be nitrogen (N) limited or NP co-limited (Axler et al., 1991a and 1994).
Representative epilimnetic samples for nutrient enrichment bioassays and APA measurements were collected using van Dorn or Niskin water samplers and composited from depths of 0.1, one and two meters in a 20 liter polyethylene container for the northeastern Minnesota lakes or from one, three, and five meters and composited in an 8 liter polyethylene container in Lake Superior (where epilimnion depth was greater). They were kept in the dark, on ice, and transported to the Natural Resources Research Institute (NRRI) laboratories in Duluth, MN within 4–6 hours. Water samples were analyzed for chlorophyll-a (chla-a), soluble reactive phosphorus (SRP), and dissolved inorganic nitrogen (DIN, ammonium and nitrite + nitrate) by standard methods (APHA 1989; Owen & Axler, 1991). Limits of detection (LOD) were calculated as approximately 2 l 1 for SRP and 5 g l 1 for ammonium and nitrite + nitrate by assuming LOD was equal to 3 the standard deviation of 10 replicates having concentrations of 20 g SRP l 1 and 50 g N l 1 for each nitrogen species (Owen & Axler, 1991). Growth bioassays Nutrient enrichment growth bioassays on northeastern Minnesota lakes using natural phytoplankton communities were performed in August and October 1990 by filling 1 liter clear polyethylene cubitainers with lake water and enriching them with either deionized (DI)water (controls), 10 mg l 1 phosphorus (KH2 PO4 -P), 100 mg l 1 nitrogen (NH4 NO3 -N), or both nitrogen and phosphorus (NP). Samples were prefiltered using 80 m netting to remove macrozooplankton. Triplicate cubitainers of each treatment were incubated outdoors in a 1000 l stock tank filled with water. Water temperatures approximated the near-surface temperature of the sampled lakes. Approximately 85% of photosynthetically active radiation (PAR) was transmitted through the cubitainer wall as measured with a Licor quantum sensor (LI192SB). APA, chl-a, and nutrient analyses were performed on initial lake water samples and on aliquots collected at the mid- and end-points of 7 to 10 day bioassays. Chl-a samples (collected by filtering approximately 100 ml aliquots through GF/C filters) were measured fluorometrically after extraction in 90% acetone (Axler & Owen, 1994).
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147 For Lake Superior bioassays, 170 ml of lake water from site LR3 or near Knife River was added to 250 ml Pyrex Erlenmeyer flasks. Stock P or DI-water (controls) were added to triplicate flasks. The flask tops were covered with small plastic petri dishes. The flasks were incubated at ambient lake temperatures in a laboratory incubator set for light saturation (PAR 200– 250 mol m 2 sec 2 ) on a 14/10 hour day/night cycle. Five ml aliquots were withdrawn every second to third day over a period of 6–10 days for in-vivo chlorophyll fluorescence using a Turner Designs model 10 fluorometer with appropriate emission and excitation filters (APHA 1989). Fluorescence values were corrected for non-chlorophyllous excitation by subtracting GF/F filtrate (< 0:7 m) fluorescence values. This method was used on the Lake Superior bioassays because it was less time and resource intensive. Alkaline phosphatase activity The fluorometric procedure of Perry (1972) was used as modified by Healey & Hendzel (1979) and Detenbeck (1987, pers. comm.) on lake and growth assay samples. Non-saturating substrate additions were used which allow greater sensitivity than methods with higher substrate additions (which have high background fluorescence). Samples were analyzed as soon as carboys arrived in the lab. APA was also measured on treatments during some bioassays. Lake water was amended with pH 8.5 Tris buffer with 200 mg Ca l 1 as CaCO3 (as per Healey & Hendzel, 1979) to a final concentration of 1 mmol l 1 . Fluorometer tubes with 4.5 ml of buffered sample were incubated at 55 C in 1990 (Detenbeck, 1987) or 35 C in 1991 (Healey & Hendzel, 1979) for 5 minutes. Healey & Hendzel (1979) and Pick (1987) found that 35–40 C was the optimum temperature for APA in a variety of lake samples. This change in the method between years would only affect comparisons between 1990 and 1991 Lake Superior measurements. Next, 0.5 ml of 1.0 mM 3O-methylfluorescein phosphate (MFP) in the pH 8.5 buffer was added to the 4.5 ml sample. The rate at which non-fluorescent MFP was cleaved into fluorescent methylfluoroscein (MF) and orthophosphate was measured for two minutes at 15 second intervals on a Sequoia-Turner 450 fluorometer with quartz halogen lamp, NB440 primary filter, and SC515 secondary filter. The fluorescence values were compared to a calibration curve of MF standards prepared for each sampling from a stock solution stored at < 0 C. Quality assurance procedures (Owen & Axler, 1991) were used
to monitor calibration curve parameters such as regression slope and intercept. The coefficient of variation (r2 ) for calibration curves exceeded 0.9999. MFP cleavage rates from unfiltered samples represented total APA, rates from 0.45 m Millipore filtered samples represented soluble APA. Differences between the two values were reported as particulate APA. For lake water samples, triplicate unfiltered and duplicate filtered sub-samples were used to determine analytical precision. During bioassays, duplicate unfiltered and single filtered samples from composite samples of each treatment were used to determine analytic precision. Background MFP breakdown and possible bacterial contamination were measured with autoclave-sterilized blanks and sample water with 1 M NaOH/0.2 M EDTA (Detenbeck, 1987). Based on algal cultures and natural phytoplankton populations, Healey & Hendzel (1979) suggested that particulate APA of < 3:0 nmol mg chl-a 1 h 1 indicated no P deficiency; 3.0–5.0 nmol mg chl-a 1 h 1 slight P deficiency; and > 5:0 nmol mg chl-a 1 h 1 severe P deficiency. Lake Superior samples were size-fractionated for APA in 1991 by measuring activity levels of filtrate passing through 20, 2.0, 0.45 and 0.2 m filters with low pressure using 60 ml hand-held syringes. Though size fractionation may not always separate bacterioplankton and phytoplankton, Cotner & Wetzel (1991) reported that a mean of 94% of chl-a was retained on 1 m pore size filters, while a mean of 95% of bacteria passed through those filters. Differences among treatment means for chl-a, nutrients, and APA measurements at each time and comparisons within treatments at different times were evaluated by Student’s t test (Steel & Torrie, 1980). Unless otherwise noted, statistical significance a should be assumed to be p < 0:05. Results Northeastern Minnesota lakes APA and chl-a measurements were made on samples from each northeastern Minnesota lake in August 1990 during the growth bioassays (see Figure 2). For Thrush Lake, only the NP treatment values were significantly higher than control values. APA increased in the control and N treatments; APA also increased in the NP treatment, although not until after day 4 when added P was depleted from the water. APA in the P
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Figure 2. Variations in chl-a (left), particulate APA (center), and SRP and DIN (right) during growth bioassays of northeastern Minnesota lakes during August, 1990. Different letters indicate differences in means at p < 0:05. Error bars are one standard deviation.
treatment was lower than other treatments on day 8. Each treatment was significantly different from other treatments and the control. In Dunnigan Lake, phytoplankton growth showed significant stimulation from N and significantly greater stimulation from NP. Our definition of co-limitation includes responses where N and P together are needed for significant algal growth
and when NP stimulates growth significantly more than N or P alone–even when N or P alone is stimulatory (Axler et al., 1994). The Dunnigan Lake bioassay’s APA levels decreased in the NP treatments, but not in the P or N treatments. APA in controls increased significantly between day 4 and day 8. In the Divide Lake bioassay, chl-a content was higher than control in
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149 the N treatments, but the NP treatment values were not statistically different from N alone. In the P alone and NP treatments. APA decreased APA decreased to levels below initial values while control values increased to values above other treatments by day 4. The APA values for the N treatment remained near initial values at the assay’s mid-point then increased. The large error associated with the day 8 N treatments (due to one extremely high value) was not significantly different than other treatments. For Sand Lake, the treatments with nitrogen had significantly higher chl-a than control by day 6. For APA, N treatments were significantly greater than controls on day 4. None of the other treatments were significantly different than control at the end of the bioassay. Uptake of added nutrients (Figure 2) during the assays generally led to depletion of SRP (except in Sand Lake) but did not deplete DIN (except for Divide Lake). There were no significant differences in uptake among the N vs NP or P vs NP treatments during these assays except for Sand Lake where P uptake was greater in the NP than P treatment during the first four days. Also, NH+ 4 -N generally was taken up at a faster rate than NO3 -N (Axler et al., 1991a). In addition to the August assays, Dunnigan Lake and Thrush Lake were assayed for APA in October 1990 growth bioassays (Figure 3). In Thrush Lake, chl-a was significantly greater in the P and NP treatments than in the controls by the end of the assay. All of the treatments had significantly lower APA values than control. Treatments with P alone and NP had lower APA than N alone treatments on day 5, but not on day 8. In Dunnigan Lake, chl-a values in the N treatments were higher than controls and NP higher than N at the end of the assay. APA values showed no significant treatment differences on day 5, but on day 8, the P and NP treatments were lower than control and N treatments. During these bioassays, added DIN was not depleted while SRP was depleted to the limits of detection. Again, there were no significant differences in DIN or SRP uptake among the N vs NP or P vs NP treatments during these assays. Lake superior Lake Superior LR3 samples showed strong growth responses to phosphorus additions (Figure 4) with the exception of the 2 November 1990 sample. The responses suggest a seasonal pattern in the magnitude of response to phosphorus additions. Particulate APA measurements for 1990–91 also indicate a pos-
sible seasonal pattern with the greatest activity found during August of both years (as shown in Figure 5). An exception to this pattern was measured on August 21, 1990 (see below). Measurements at other sites (Figure 5, lower left) show total APA values which were not significantly different than LR3 APA values with the exception of seasonal extremes (8 May 1991 and 9 October 1991). Total APA differences at LR3 during 1990 and 1991 were not significantly different when compared by month (even at different sites) with a single exception, the 9 October 1991 sample (p > 0:01) which was the only one taken during unstratified conditions (Axler et al., unpublished data). Dissolved activity was an important component of APA, at all Lake Superior sites, accounting for 33–79% (mean 54%) of the total activity. Dissolved APA was generally higher in Lake Superior than in the northeastern Minnesota lakes (mean 26% of total activity). In 1991, several size-fractionation experiments revealed that at LR3 in August, most of the particulate enzyme activity was associated with < 2:0 m particles. The Gooseberry River sample was also dominated by particulate activity. At the Knife River site, particulate activity was dominated by the > 20 m size fraction. As a result of a strong wind storm from August 18–20, 1990, suspended particulates increased turbidity of a 100 km2 area of Lake Superior near Duluth. The ‘clear’ water nearer to shore (LR1 and LR2) was visually obvious and demonstrated that the off-shore turbidity was not derived from North Shore streams. Particulate APA were lower (at p < 0:05) at LR3 than at the other non-turbid or ‘clear water’ sites along the Lester River transect or at LR3 on 21 August (Figure 5, lower right). APA levels in Lake Superior were normalized to chl-a and particulate organic carbon (POC) (Table 1). Most of these normalized levels are above Healey & Hendzel’s (1979) criteria for severe phosphorus deficiency. Exceptions were from 8 May and 9 October 1991 samples and the sample associated with the particle suspension event. Size fractionated chl-a data were also available on 5 September 1991. The > 20 m fraction indicated high particulate APA but also had a high standard deviation (Table 1). In July, 1992, APA kinetics were measured on samples from near Knife River (Figure 6). From these data, the Michaelis-Menten equation (see Discussion) was used to estimate the maximum rate of hydrolysis (Vmax ) and half saturation constants (Km ) as per Healey & Hendzel (1979). Maximum soluble APA was 61% of maximum total APA at that time, and substrate lev-
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150
Figure 3. Variations in chl-a (left) and particulate APA (right) during growth bioassays of northeastern Minnesota lakes during October, 1990. Different letters indicate differences in means at p < 0:05. Error bars are one standard deviation.
Figure 4. Phosphorus addition growth bioassay responses of Lake Superior phytoplankton as measured by in vivo chlorophyll florescence for LR3 samples collected at indicated dates in 1990. The 19 September sample was collected from Knife River. All treatment means were different than control values at p < 0:05.
els of 2 mmol l 1 did not saturate the substrate (mean measured total APA value was 82% of estimated value).
Discussion Northeastern Minnesota lakes (N, NP co-limited) A variety of methods have been used to determine APA in algal cultures and natural systems (Petterson & Jansson, 1978; Jansson et al., 1988). The method used in the present study was not most similar to that of Healey & Hendzel (1979). Their method used higher
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151
Figure 5. APA during 1990 (upper) and 1991 (middle) at Lester River site LR3; and at other Lake Superior sites (lower left): GR-Gooseberry River sampling site, KR-Knife River sampling site. The lower right panel represents particulate APA along the Lester River sampling on 21 August 1990 (after storm). Station LR-5 as a large patch of clear water at the time of sampling. Error bars are one standard deviation of the total APA. Different letters indicate differences in means at p < 0:05.
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Figure 6. APA as a function of MFP additions for Lake Superior Knife River station on 14 July 1992. estimated as in Healey and Hendzel (1979) using a Hanes regression.
substrate concentrations (10 mmol l 1 MFP) which reduces assay sensitivity at low activity levels (due to increased background signal interference), but yields estimates near the maximum rate of hydrolysis (Vmax ). When the Michaelis-Menten equation
V
=
Vmax S=(S + Km)
is solved with our substrate (S ) additions of 0.1 mmol l 1 and Healey & Hendzel’s (1979) reported half saturation constants (Km ) of 0.2–4.0 mmol l 1 and Vmax of 1.4–440 mmol l 1 h 1 from algal cultures and natural phytoplankton populations (Healey & Hendzel, 1979; Pick, 1987), the results indicate that our rates were 0.1 to 0.025 of Vmax . However, because the substrate additions used were constant in these experiments, the patterns found in these experiments would likely be similar if saturating substrate levels were used. Even without correcting for potential underestimation due to substrate differences, Healey & Hendzel’s (1979)
Km and Vmax ( standard error) were
criteria for P deficiency (i.e. > 30 nmol MF mg chla 1 h 1) was exceeded in most of the growth bioassay experiments (Figures 2 and 3). Variations in APA during the growth assay suggest interaction of many factors (Figure 2 and 3). APA in samples receiving P alone were significantly lowere than the control APA values in all cases except Dunnigan Lake (summer and fall) by the bioassay mid-point. The Dunnigan Lake results suggests that stored P in algal cells, regeneration of P by bacteria or microzooplankton, or non-nutrient factors may be important in those samples. The N alone treatments had significantly higher APA than P alone on four of six occasions. APA in NP enriched samples was lower than control values at the mid-point, then increased (though not always significantly) by the end-point (with the exception of Sand Lake during the summer bioassay). Those results suggest that when N is available to these algal cells, they sequestered available extracellular P, lead-
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153 Table 1. Particulate alkaline phosphatase activity normalized to chl-a and particulate organic carbon (POC). Values are means one standard deviation. NA means POC was not available
nmol MF g chl-a 1 h 1
Date
Site
6/29/90 7/17/90 8/15/90 8/21/90
LR3 LR3 LR3 LR1 (clear) LR2 (clear) LR3 (turbid) LR5 (clear) LR3
2:1 8:1 7:8 11:2 13:9 4:4 13:3 6:6
Gooseberry Knife LR3 Knife > 20 m 2–20 m 0.2–2 m LR3
4:3 3:5 8:2 1:9 5:8 0:9 5:9 0:6 78 57 4:0 0:7 2:2 0:2 0:2 0:2
10/10/90 6/12/91 7/25/91 8/12/91 9/5/91
10/9/91
0:0 3:4 3:2 1:7 2:0 0:8 1:3 2:8
nm MF g POC 1 h 1 NA 22:7 19:7 NA NA 11:0 57:5 14:2 21:4 20:5 24:4 NA NA NA NA NA
9:4 8:0 3:5 13:5 6:1 17:3 5:4 3:3
Table 2. Comparison of nutrient limitation predictions at end of bioassays based on biomass (chl-a) and particulate APA data Lake
Month
Biomass prediction
APA prediction
Thrush Dunnigan Divide Sand Thrush Dunnigan
August August August August October October
NP N and NP N N P N and NP
P not P P not P P P
ing to increased P demand (indicating N or NP colimitation). Examples of this include Thrush Lake in August and Dunnigan Lake in October. Control treatments generally increased as might be expected in an enclosed container where P sources to phytoplankton in the epilimnion of a lake (such as macrozooplankton recycling or atmospheric inputs) are eliminated. APA changes in the six growth assays were generally consistent with nutrient enrichment growth assay results (as shown in Table 2). Our criteria for predicting P growth limitation from APA changes was that if APA in P and NP treatments decreased while control and N alone treatments increased, some form of P limitation
was possible. All other results were considered to predict ‘not-P’ limitation. Surprisingly, Divide Lake in August 1990 had initial APA values indicating severe P deficiency, but N deficiency was clearly indicated by growth bioassay and by ammonium enhancement assay (Axler et al., 1994). P addition caused only a slight chl-a enhancement over NP, relative to N alone. Four of six APA predictions agreed with biomassbased predictions. These results indicate that enzyme activity is repressed by the addition of P and support the conclusion that P availability can control APA levels. Also, a comparison of initial APA levels and growth bioasay results from the northeastern Minnesota lakes during 1990–1991 shows that using APA as a physiological indicator predicted P or NP limitation correctly in 12 out of 14 outcomes (Axler et al., 1994). However, without additional information regarding the nitrogen status of phytoplankton communities (which was gathered in the 14 bioassays), important nutrient limitation information could be missed. Elser & Kimmel (1986) measured APA changes of phytoplankton following P enrichment and found decreased activity rates of the enzyme after 24 hours when activity was high, but no significant inhibition when activity was low. APA sometimes increased in their experiments (relative to initial values) after about 4 days even in samples receiving 10 mg P l 1 . Pick (1987) also found significant APA inhibition from added P in as little as three hours, but much longer in most other instances. These experiments demonstrate that during bioassays, the APA levels are often affected by N additions as well as P. Lake Superior (P-limited) The APA was used as an indicatior of P-deficiency, to assess seasonal changes in phytoplankton P status, and to help assess spatial variation in phytoplankton Pmetabolism due to a severe wind storm and resuspension event. The seasonal pattern found using this physiological assay corresponded well with results from bioassays and the storm studies indicating that it is a useful tool for assessing the phosphorus limitation status of Lake Superior phytoplankton. The particulate APA normalized to chl-a values (Table 1) indicated P-limitation for nearly all samples (i.e. > 30 nmol MF mg chl-a 1 h 1 [Healey & Hendzel, 1979]) despite the non-saturating substrate levels used. This was consistent with growth bioassay results and the presence of dissolved inorganic-N at levels > 250 g l 1 throughout the growing season
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154 (Axler, unpublished data). Silicon values were also relatively high, about 2 mg Si l 1 with almost no seasonal or depth variation, and so P-limitation alone would be the expected nutrient response. APA data from a station off the eastern end of Isle Royale for 1990–91 using saturating MFP concentrations suggest that Vmax may be 8–10 higher than our APA measurements (Guildford et al., 1994). The Isle Royale and LR3 stations had similar POC and chl-a values. Analysis of a kinetics experiment using water collected from the Knife River site in July 1992 suggested that our Vmax values could have been underestimated by a factor of 5 (Figure 6). Therefore, our data indicates severe P stress. Particulate organic matter (POM) is also used to normalize APA for indicating P limitation. Healey & Hendzel (1979) and Petterson (1980) reported Pdeficiency criteria for particulate APA of > 46 and 76 nmol MF mg POC 1 h 1 , respectively. Our values exceed either of these criteria (assuming POC 0:43 POM [Healey, 1973]) or Istvanovics’ et al. (1992) extreme P deficiency criteria of 32 nmol mg chla 1 h 1 total APA if corrected for substrate limitation as described above. Pick (1987) found that phosphate additions greater than observed under natural conditions in Lake Ontario (33 g P l 1 ) were needed to inhibit APA to 50% of its initial level within 18 hours. Her conclusion was that APA may not be a sensitive indicator of phytoplankton P deficiency. However, the APA levels observed by Pick in Lake Ontario were at levels associated with P deficiency and indicated spatial and temporal changes due to water movements. We have several data sets showing APA reductions of > 50% after 3–4 days in response to additions of 10 g P l 1 (Figure 2). There are several possible reasons for lower levels of APA associated with the turbidity caused by the wind storm (Figure 5). Turbid water reduced Secchi depth to 0.6 m and mid-day light at 2 m to 8.9 mol m 2 s 1 , a level shown to limit phytoplankton APA (Wynne and Rhee, 1988). SRP remained below detection levels at LR3, suggesting that it should not have inhibited APA. However, total phosphorus greatly increased from 5 to 30 g l 1 . Primary productivity of surface phytoplankton increased compared to the other transect stations, and these higher rates were sustained for nearly a week (Axler et al., 1991b). These results suggest that the phytoplankton were able to utilize some of the total phosphorus (presumably associated with the particles) either directly, or indirectly due to APA of algae or bacteria attached to large particles, or via P-mineralization (Engle & Sarnelle, 1990). Low light and relatively high
phosphorus utilization both may have been responsible for the lower APA levels at LR3 on 21 August 1990. Several hypotheses could explain the high proportion of dissolved phosphatase activity in natural systems. Culture studies of algal and bacterial cells have shown that algal cells are likely to have extracellular phosphatase release (see summary by Jansson et al., 1988). Cell damage (via zooplankton grazing, filtration, or death) may release phosphatases and zooplankton digestive enzymes may also be an APA source (Wynne & Gophen, 1981). Once released from the cell surface, alkaline phosphatases have a half life of days to weeks (Jansson et al., 1988). The proportions of APA measured in the size fraction experiments are within the ranges reported in the literature. Pick (1987) found most particulate APA in Lake Ontario associated with < 5 m particles, though she reported one sample with 40% of APA in the > 12 m fraction. Berman et al. (1990) measured a wide range (< 5 to 70%) of > 20 m particulate activity in oligotrophic Lake Kinneret, Israel. These differences may be due in part to the relative ability of different algal taxonomic groups to produce phosphatases (Elser et al., 1986). The chl-a normalized APA levels associated with the > 20 m fraction was high in Lake Superior (Figure 5) suggesting that algal cells in this size fraction may be under greater P stress than smaller fractions. Suttle et al. (1991) presented evidence suggesting that algal cells > 3 m were outcompeted for P by smaller cells; the larger cells were under greater P stress. This may indicate the need to study competition between bacteria and algae in these systems. Another factor that may play a role in phosphorus regeneration in our study lakes is bacterial cell-surface 50 -nucleotidase (5PN). The activity of this enzyme is not inhibited by orthophosphate at concentrations found in aquatic environments (Ammerman & Azam, 1985). Cotner & Wetzel (1991) studied this enzyme in Lake Michigan and in a smaller, more productive lake. Their results suggest that 5PN activity may be a more important P-pathway than APA in oligotrophic as opposed to eutrophic systems. They hypothesized that because bacteria are likely to be limited by organic carbon in oligotrophic systems, their enzymes (5PN) may liberate orthophosphate, which would then be available to P limited algae. In eutrophic habitats, bacteria would take up a greater proportion of enzyme-liberated phosphate. A comparison of this pathway in Lake Superior (P-limited) and Thrush Lake (which switches from N or NP to P-limitation seasonally) may contribute sup-
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155 porting evidence for this hypothesis as well as a better understanding of phosphorus pathways.
Conclusions Our results suggest that caution should be used when interpreting APA data. The growth bioassay experiments performed for the northeastern Minnesota lakes suggest that when co-limitation or secondary N limitation is present, APA may indicate P deficiency without a corresponding growth increase from P additions. APA increases after P uptake probably indicate that the phosphorus was utilized rather than simply stored. Lean et al. (1989) suggested that physiological assays such as APA may be more useful as indicators of nutrients ‘not’ limited to natural populations. If this type of interpretation is used on these results, seldom could P be discounted as a limiting nutrient. The strength of APA seems to be as supportive evidence in multiple assay assessments and in assessing short term temporal or spatial changes in phytoplankton nutrient status.
Acknowledgements J. Ameel, C. Host, C. Owen, C. Radosevich, S. Randall, and C. Tikkanen contributed to the laboratory portion of this study. R. Strassman and E. Swain from MPCA, D. Wright, M. Danks, W. Popp, and P. Eiler from MDNR, and R. Hicks, P. Aas, D. Pasco, and D. Weaver of UMD collected many of the field samples. R. Hicks also supplied POC data for Lake Superior on 21 August 1990. E. Markham, T. Majerle, and A. Bellamy assisted with administration or report preparation. Anonymous reviewers helped to improve this manuscript. Funding was contributed by the Minnesota Pollution Control Agency, the Minnesota Department of Natural Resources and the Graduate School of the University of Minnesota. This is Contribution # 212 from the Center for Water and Environment at the Natural Resources Research Institute.
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