Acidic Precipitation and Its Consequences for Aquatic Ecosystems: A Review

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Acidic Precipitation and ItsConsequences for Aquatic Ecosystems:A ReviewTerry A. Haines aa United States Fish and Wildlife Service, Columbia NationalFishery Research Laboratory, Field Research Station, Universityof Maine, Orono, Maine, 04469, USAPublished online: 09 Jan 2011.

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TRANSACTIONS of the

AMERICAN

FISHERIES SOCIETY

November 1981

VOLUME 1 10

NUMBER 6

7¾ansactions of the American Fisheries Society 110:669-707, 1981

Acidic Precipitation and Its Consequences for Aquatic Ecosystems: A Review

TERRY A. H^INES

United States Fish and Wildlife Service, Columbia National Fishery Research Laboratory Field Research Station, University of Maine, Orono, Maine 04469

Abstract

Precipitation in Europe and eastern North America has become acidic, a result of increases in sulfuric and nitric acid aerosols produced by fossil-fad combustion, metal smelting, and industrial processes. The increased use of tall smoke stacks and particle removers has increased long-range transport of acidic gases. Some metals and organic compounds also are transported atmospherically and deposited in acidic precipitation. In regions where acid-neutralizing capacity of soils and water is low, the pH of lakes and streams has decreased and concentrations of metals have increased. Aquatic organisms have hccn affected in all trophic levels (decomposers, primary producers, primary and secondary consumers); abundance, production, and growth have hccn reduced and sensitive species have been lost. Fish have suffered acute mortality, reduced growth, skeletal dcœormitics, and especially reproductive failure. Valuable commercial and recreational fishcries have hccn lost in certain areas and such losses will become more widespread if acidic precipitation continues. Remedial or mitigative actions directed toward the problem include hatchery production of acid-tolerant fish and chemical neutralization of selected lakes and streams. The ultimate solution is reduction of the sources of atmospheric acid.

Contents

Chemistry of Precipitation ..................... 670 Sulfates and Nitrates .......................... 670 Organics and Metals .......................... 673

Chemistry of Surface Waters ................... 675 Changes in pH and Alkalinity ................... 675 Organics and Metals .......................... 677 Seasonal Patterns ............................. 679

Changes in Non-Fish Aquatic Biota ............. 679 Bacteria and Fungi ........................... 679 Primary Producers ............................ 680 Invertebrates ................................. 681

Amphibians .................................. 684 Changes in Fish Populations .................... 684

Mortality .................................... 685 Reproduction ................................. 687 Growth ...................................... 690

Skeletal Deformity ............................. 690 Metal Uptake ................................ 690

Ecosystem Effects ............................. 691

669

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670 HAINES

Vulnerability of Surface Waters to Acidification .. 691 Remedial Action .............................. 692

Neutralization ................................ 693

Selective Breeding ............................. 694 Maintenance Stocking ......................... 695 Other Actions ................................. 695

Summary ..................................... 695 Acknowledgments ............................. 695 References .................................... 695

The term "acid rain" has been coined to de-

scribe precipitation that has a pH lower than 5.6 (the approximate pH of distilled water equilibrated with atmospheric concentrations of COs). Historical data on precipitation chemistry suggest that in rural areas remote from sources of acidic gases the pH rain and snow was generally 5.0 or higher. Large regions in Europe and North America now receive precipitation below pH 4.7, and adverse effects on aquatic biota have been observed in areas where surface waters have been acidified. In this re-

view, I summarize current understanding of this phenomenon, its adverse effects on aquatic ecosystems, and possible mechanisms for deal- ing with the problem.

Chemistry of Precipitation Acidic precipitation was first described in En-

gland by Gorham (1955) and was first recog- nized as a regional phenomenon in Scandinavia in the late 1960s. The European Atmospheric Chemistry Network, operated by the Interna- tional Meteorological Institute, Stockholm, Sweden, began operation in Sweden in 1946 and expanded to most of western Europe by 1955 (Whelpdale 1979). Increasing acidity of precipitation in Europe, as documented by this network, was reported by the Swedish govern- ment in 1967 (Oden 1976). The weighted mean annual pH of precipitation in southern Nor- way, for example, declined from 5.0-5.5 in the late 1950s to 4.2-4.4 in the mid-1970s (Dovland et al. 1976).

Sulfates and Nitrates Although the pH of precipitation can be af-

fected by natural sources of acidic gases, such as hydrogen sulfide (HsS) and sulfur dioxide (SO2) from volcanos or sulfur bacteria, there are no apparent trends in natural phenomena that account for the increasing acidity of precipitation. Delmas et al. (1980) found that Ant-

arctic snow had a pH of 5.2 to 5.5 and had been constant for thousands of years. Periods of vol- canic activity produced only temporary pH declines.

Back trajectories of storms that deliver large amounts of acid to Scandinavia indicate that the

storms pass over heavily industrialized areas of central and western Europe and England. In- creased acidity of precipitation is attributed to production of SOs and nitrogen oxides (NOx) from fossil-fuel combustion, which were trans- formed into sulfuric and nitric acids and trans-

ported long distances in the atmosphere. Fossil- fuel combustion in Europe has increased, and the sulfate and nitrate content of precipitation has increased as pH has decreased (Dovland et al. 1976). Acidity of precipitation in Scandina- via is correlated with air-mass trajectories and emission sources of SO2 (Nordo 1976).

Precipitation is now more acidic in eastern North America than in Scandinavia. The me-

dian pH in 1978-1979 ranged from 4.0 to 4.4 in the northeastern United States and south-

eastern Canada (Atmospheric Environment Service 1979; National Atmospheric Deposition Program 1980). Major sources of SOsand NOx emission are in the region from Pennsylvania to Indiana, and prevailing winds and storm tracks carry these emissions to the northeast, as shown in Figs. 1-3 (Cogbill and Likens 1974; Schlesinger et al. 1974; Altshuller and McBean 1979). There are no long-term precipitation chemistry networks in North America like those in Europe. Cogbill's (1976) precipitation pH maps for the United States (Fig. 4) are based on networks that operated in 1965-1966 and 1972-1973, during which time pH was measured directly, and in 1955-1956, when pH was calculated. Precipitation already was acidic in 1955-1956, and has become steadily more so since. Cogbill (1976) interpreted indirect evidence, such as the presence of methyl orange alkalinity, to mean that precipitation in New

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ACIDIC PRECIPITATION: DEGRADATION OF AQUATIC ECOSYSTEMS 671

FIGURE 1.•Average annual sulfur dioxide emissions in eastern North America, 1973-1978. Data are from Environment Canada and the United States SURE H data base (Altshuller and McBean 1979).

York, Virginia, and Tennessee during the 1920s was low in acidity--probably pH exceeded 5.5.

In the United States, the estimated annual

SO.• emissions have increased only slightly between 1950 and 1975 (from 22.0 to 25.7 million tonnes), but NOx emissions have nearly tripled (8.1 to 22.2 million tonnes). Sulfate concentrations measured in precipitation during 1955-

1956 (Junge and Werby 1958) were higher than those reported for 1978-1979 (National At- mospheric Deposition Program 1980)--2.0-3.0 mg/liter versus 1.5-2.0. Likens and Bormann (1974) reported a similar decline in sulfate in precipitation at the Hubbard Brook Experi- mental Forest in New Hampshire, and an even larger decline from 1915-1920 to 1970 in New York. However, nitrate in precipitation in-

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672

F•CVRE 2.•Average annual nitrogen oxide emissions in eastern North America, 1973-1978. Data are from Environment Canada and the United States SURE H data base (Altshuller and McBean 1979).

creased sharply at Hubbard Brook from 1965 to 1973, and in New York from 1915 to 1970.

The cause of acidic precipitation thus is not as apparent in the United States as in Europe. The discrepancy between sulfate cot•centration and acidity of precipitation may be due to changes in emission chemistry; before 1950 much of the sulfate may have been in un-ionized particles or neutralized by bases from smoke particles

(Likens and Bormann 1974). This explanation is consistent with the shift from coal to other

fuels that begat• in the early 1900s (Singer 1970) and with the increased use of particle precipitators in smokestacks since about 1950 (Patrick et al. 1981). The acid portion of precipitation in eastern North America is roughly two-thirds sulfuric acid and one-third nitric

acid, plus trace amounts of hydrochloric acid.

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Depression Tracks

Prevailing Winds

/

FIGURE 3.•eneralized weather patterns for North America, after Kendrew (1953) and Barrett (1974).

Although organic acids are present, they do not contribute appreciably to the free hydrogen ion content (Galloway et al. 1976; Marsh 1978). The increased acidity of precipitation since 1950 may be the result of increased contribu- tion from nitric acid.

On a global basis, natural sources of sulfur emissions to the atmosphere probably exceed anthropogenic emissions from fossil-fuel combustion and metal smelting (Galloway and Whelpdale 1980). However, the anthropogenic sources are concentrated in the north temper- ate latitudes; one-sixth of all anthropogenic sulfur produced in the world comes from an area in the northeastern United States covering only 0.6% of the global land area (Shinn and Lynn 1979). Analysis of sulfur content in ice cores from Greenland indicated that sulfur deposition is increasing, probably a result of increasing anthropogenic emission (Herron et al. 1977). Hofmann and Rosen (1980) reported that small aerosol sulfate particles in the strato- sphere have increased 9% per year for 20 years.

The increasing use of tall smokestacks to prevent local concentrations of air pollutants un- doubtedly contributes to increased deposition of pollutants in remote areas (Miller 1975; Pat- rick et al. 1981).

Organics and Metals

Pollutants other than acids are transported in the atmosphere and deposited with precipitation. The complex organic compounds in precipitation are potentially important but poorly understood (Gorham 1976). In Norway, more than 450 organic compounds have been detected in precipitation (Lunde et al. 1977). Organic compounds detected that are presumably of anthropogenic origin include alkanes, polycyclic aromatic hydrocarbons (PAHs), phthalic acid esters, fatty acid ethylesters, pesticides, and a diverse group of common industrial chemicals such as polychlorinated biphenyls (PCBs), benzaldehyde, tri-n-butyl- phosphate, and diphenylamine (Cohen and Pinkerton 1966; Lunde and Bjorseth 1977;

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674 uAi•Es

5.0

. .,.•:- •--• 5.6 • - ß

/•._.•. 5.6

1955-1956 1965-1966

' 4.4

i . .. ( ½ '.

1972- 1975 1979

FIGURE 4.•hanges in mean annual pH of precipitation, 1955-1956 to 1979, over eastern United States. Data for 1955-1973 are after Cogbill (1976). Data for 1979 are from National Atmospheric Deposition Program (J. Gibson, Colorado State University, personal communication).

Lunde et al. 1977; Murphy and Rzeszutko 1978; Schrimpff et al. 1979; Strachan and Hu- neault 1979; Alfheim et al. 1980). These compounds are either produced by fuel combustion or are introduced into the atmosphere by combustion processes, and thus are likely to be related to acidic precipitation.

Fossil fuels (especially coal) contain trace elements, some of which are vaporized during

combustion. For example, gaseous selenium emissions from coal combustion are estimated

to contribute 1.5 to 2.5 times more selenium to

the environment than natural weathering (An- dren et al. 1975). Block and Dams (1975) reported that halogens, mercury, selenium, arsenic, and antimony are preferentially volatized during coal combustion. More than 90% of the mercury in coal may be vaporized (Billings et

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ACIDIC PRECIPITATION: DEGRADATION OF AQUATIC ECOSYSTEMS

TABLE 1 .--Metals in rain and snow from areas receiving nonacidic (pH > 5.6) and acidic precipitation.

675

Metal Type of precipitation, area

(reference) a AI Cd Cu Fe Hg Mn Ni Pb Zn

Concentration in precipitation (•g/liter) Nonacidic

Snow, Manitoba (1) < 1 <2 Rain and snow, Nebraska (2) 350 0.3 4 Rain, north Norway (3) 0.4 Annual mean, north Norway (4) 0.2

Acid ic

Annual mean, south Norway (3) 0.3 5.3 Annual mean, Sudbury (1) 0.6 11 36 Rain and snow, Minnesota (5) 56 0.2 4 38 Rain, New Hampshire (6) 0.6 Annual mean, Massachusetts (7) 49 <0.5 27 Annual mean, central Europe (8) 190 1.2 21 150 Snow, south Sweden (9) 7 9 216 Snow, south Norway (10) 0.5 2 Snow, Quebec (11)

<1 <2 <1 <1

5 5 10

4 3 38

4 4 8

Nonacidic

North-central United States, five stations (12)

Acidic

Northeastern United States, five stations (12)

New Hampshire (6) Central Europe (8) New York (13)

7 13 29

30 7 21 26 3 I 7 93

0.06 13

3 <5 26 <75

0.04 20 I 38 370

5 48 64 311

3 28

2.9

Annual deposition from precipitation (g/hectare)

5 7 26 7 22 58

326 293 190 60 703 1,154 9 0.8 196

2,000 13 224 1,600 0.4 200 15 405 3,900 173 183 235

a References: (1) Beanfish and Van Loon 1977; (2) Struempler 1976; (3) Semb 1978; (4) Hanssen et al. 1980; (5) Eisenreich et al. 1980; (6) Schlesinger et al. 1974; (7) Environmental Measurements Laboratory 1980; (8) Heinricks and Mayer 1977; (9) Dickson 1975; (10) Wright and Dovland 1977; (11) Delisle et al. 1979; (12) Lazrus et al. 1970; (13) Troutman and Peters 1980.

al. 1973). It is estimated that 99% of the lead in the atmosphere is of anthropogenic origin (Elias and Patterson 1980).

Metal contents of precipitation are higher in areas where precipitation is acidic than else- where (Beamish and Van Loon 1977; Semb 1978; Hanssen et al. 1980). At least eight metals show this trend: lead, zinc, copper, iron, manganese, nickel, mercury, and cadmium (Table 1). Metal concentrations are highest in precipitation near smelters, urban areas, and high- ways (Beamish and Van Loon 1977; Franzin et al. 1979; Schrimpff et al. 1979), but also have been detected in remote areas (Schlesinger et al. 1974; Semb 1978). Precipitation in remote areas of Quebec and Ontario contains mercury at concentrations an order of magnitude higher than in surface waters of those areas (Delisle et al. 1979). Although mercury in the air may re- suit from natural sources, increased acidity of precipitation results in increased scavenging of

mercury from the air (Tomlinson et al. 1980). Concentrations of lead and zinc have increased

in recent Greenland ice layers over historical levels (Murozumi et al. 1969; Herron et al. 1977); however, earlier reports of similar changes for mercury have been disproved (Dickson 1972; Carr and Wilkniss 1973; Weiss et al. 1975; Appelquist et al. 1978). Neverthe- less, it appears certain that anthropogenic ac- tivities have increased the aerial transport and deposition of metals in aquatic environments hundreds of kilometers from the sources.

Chemistry of Surface Waters

Changes in pH and Alkalinity The extent to which acidic precipitation can

change an aquatic system is determined by the geochemistry, geomorphology, and hydrody- namics of the system. These determine the capacity of soil and water to neutralize acids, resist

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676 HAINES

pH change in surface waters, and absorb metals and organic compounds. In areas where the buffering capacity is low, surface-water alkalinity has been reduced or eliminated and pH has declined as a result of acidic precipitation. Such changes have been recorded in Scandinavia; Nova Scotia, Ontario, and Quebec in Canada; and Maine, New Hampshire, New York, New Jersey, North Carolina, and Florida in the United States.

Comparisons of current water-chemistry data with historical data must be made with caution.

Prior to about 1955, pH could only be measured by colorimetric methods. Colorimetric methods generally overestimate true pH, especially at low alkalinities (Likens 1975; Boyd 1977, 1980; Pfeiffer and Festa 1980). Likewise, older alkalinity data are largely fixed-endpoint titrations, which overestimate true alkalinity in soft waters (American Public Health Associa- tion et al. 1975); unfortunately the degree of error cannot be determined. Although Hen- drey et al. (1980) applied a correction to fixed- endpoint alkalinity data before making comparisons, Jeffries and Zimmerman (1980) point out that the differences are nearly random and cannot be corrected by application of a uniform adjustment. However, such an adjustment will produce a more conservative estimate of change in alkalinity over time than would be the case if no such adjustment was made.

Nevertheless, useful comparisons of recent and historical water chemistry data may be made. If recent measurements are made by the same methodology and samples are collected at the same locations, seasons, and times of day, relative differences between historical and re-

cent pH values should indicate a real change in pH. Only comparisons that use such an approach are included in this review.

In a group of lakes in Norway that were surveyed in 1933-1941 and resurveyed in 1971- 1975, the pH had declined by 0.8 to 1.8 units during the intervening period (Gjessing et al. 1976). Seip and Tollan (1978) concluded that poorly buffered surface waters in Scandinavia, Ontario, and New York had declined by 1 to 2 pH units during the last 30-40 years. Wright and Henriksen (1978), who surveyed 155 lakes in southern Norway, showed that pH was usually less than 5.5 in lakes in granitic terrain that received acidic precipitation. The mean pH in 50 lakes in southeastern Norway declined from

7.0 in the 1950s to 6.8 about 10 years later. In soft-water lakes (below 17.9 mg/liter as CaCOs), pH declined from 6.6 to 6.3, whereas lakes with harder water were unaffected (J. Okland 1980a). Lakes in Sweden have declined by as much as 1.8 pH units since the 1930s (Almer et al. 1974). Of 321 lakes on the west coast of Sweden that were examined during 1968- 1970, pH was less than 5.5 in 93%, and 4.0-4.5 in 53% (Dickson 1975; Wright and Gjessing 1976).

In Canada, Watt et al. (1979), who resurveyed 19 lakes in Nova Scotia originally surveyed in 1955, found that pH had declined in all of them, apparently from atmospheric inputs of acid. Comparison of historical and recent data for 14 Nova Scotia rivers showed a

decline in pH in all (Altshuller and McBean 1979; Thompson et al. 1980). Lakes in the vicinity of a metal smelter near Sudbury, Ontario, have been drastically acidified. Beamish and Harvey (1972) surveyed 46 lakes for which historical pH measurements were available in, and east and north of, the LaCloche Mountains. The decline in pH averaged 0.16 unit/year for lakes in and east of the mountains, and 0.08 unit/year for lakes to the north. The pH in one (Lumsden Lake) declined from 6.8 in 1969 to 4.4 in 1971, representing a 200-fold increase in hydrogen ion concentration in 10 years. Lakes in this region are much more acidic than lakes from the Experimental Lakes Area of northwest Ontario, which are in similar geological terrain but are not receiving acidic precipitation (Beamish 1976). Lakes in south-central Ontario receive acidic precipitation and in a few the pH is as low as 5.0. There is evidence of a decline

in buffering capacity in some lakes (Dillon et al. 1978; Scheider et al. 1979). Jones et al. (1980) compared pH of lakes in Laurentides Park, Quebec, in 1979 and 1980 to data from 1938- 1941 and concluded pH had declined signifi- cantly.

In 217 lakes above 610 m elevation in Adi-

rondack Park, New York, surveyed in 1975, the pH in 51% was less than 5.0; in 40 of these lakes that had been surveyed during 1929- 1937, pH was this low in only 4% (Schofield 1976). In New Jersey, recent evidence indicates that stream pH declined 0.4 unit and hydrogen ion increased 50 /zeq/liter from 1958 to 1979 (Johnson 1979). Davis et al. (1978) reported that the mean pH of large, lowland lakes in

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Maine declined from 6.18 in 1937-1943 to 6.09

in 1969-1974, a difference that corresponds to a 5.2-fold increase in hydrogen ion concentration. Two groups of poorly buffered lakes in Florida were compared by Crisman et al. (1980). In a group in north Florida that received acid precipitation, pH in a few had declined from that recorded 10 years earlier; the group in south Florida, which did not receive acid precipitation, showed no decline in pH.

Hendrey et al. (1980) compared current pH and alkalinity with that measured in 1960-1964 for 35 North Carolina streams. The mean pH declined from 6.77 to 6.51, and a decline occurred in 79% of the streams. Mean alkalinity declined from 116 tzeq/liter to 80 tzeq/liter, and 71% of the 1979 values were lower than those for 1960-1964. Similarly, in New Hampshire the mean pH in 43 lakes and streams declined from 6.66 in the late 1930s to 6.12 in the late

1970s, and 90% of the values were lower in the 1970s than in the 1930s.

In contrast, evidence of acidification of surface waters was not detected in studies in Penn-

sylvania and Wisconsin. Arnold et al. (1980) reviewed existing water-chemistry data for Pennsylvania and compared older with more recent measurements. They did not include areas affected by acid mine drainage. Of 314 usable comparisons, 34% showed reductions in pH or alkalinity or both. Although the authors interpreted these reductions as significant evidence of acidification, random variation is an

equally plausible explanation. Bush (1980) and Lillie and Mason (1980), who surveyed lakes in soft-water areas in Wisconsin, found no evidence of acidification, although many lakes had a pH below 6 and were believed to be vulnerable. Precipitation in Wisconsin has a pH above 4.4, whereas it is below 4.4 in northeastern states.

In the absence of historical data other evi-

dence may be used to demonstrate changes in surface-water chemistry related to acidification. In areas receiving neither acidic precipitation nor marine salts in aerosols, calcium and magnesium should be derived solely from weathering of rocks by H2CO3. Thus, calcium plus magnesium should be electrochemically equivalent to alkalinity. This relationship has been confirmed for remote lakes in northern Sweden

and Canada. Acidic precipitation leaches additional calcium and magnesium from watersheds

and excess hydrogen ions reduce alkalinity. Lakes in such watersheds have calcium plus magnesium equivalencies greater than that of alkalinity; excess cations are balanced by nonmarine sulfate and nitrate (Almer et al. 1978; Dillon et al. 1980; Dickson 1980; Henriksen 1980). Based on this relationship, a nomograph was prepared that successfully predicted lake pH from nonmarine calcium plus magnesium, and either precipitation pH or the nonmarine sulfate in precipitation (Wright et al. 1980; Henriksen 1980).

Organics and Metals Complex organic compounds that are pro-

duced by fuel combustion or are introduced into the atmosphere by combustion processes have been detected in lakes, remote from any known source of these substances, that also receive acidic precipitation. In Norway, PCBs have been detected in the water of remote lakes

(Alfheim et al. 1978), and PAHs have been detected in the sediments of rural and remote

lakes in the United States (Heit and Tan 1979). In two acidic lakes in the Adirondack Moun-

tains of New York, PAH compounds were found in the upper sediment layers but not in sediment deposited before about 1950. This timing agrees well with the onset of acidic precipitation (Heit et al. 1981). Atmospheric input is believed to be a major source of the PCBs in lakes Superior, Michigan, Huron, and Erie (Anonymous 1976; Murphy and Rzeszutko 1978; Eisenreich and Hallod 1979; Josephson 1979; Eisenreich et al. 1981; Murphy et al. 1981), and in Norwegian lakes (Alfheim et al. 1978).

The trace-metal content of surface waters is

affected by direct input of metals from precipitation and by uptake and leaching of metals from watershed soils and sediments. Lakes that

are acidified generally have a much higher metal content than similar lakes that are not

(Table 2). Lakes in Norway, Sweden, and Ontario have

higher concentrations of zinc, lead, copper, cadmium, and nickel than do unacidified lakes from nearby areas (Beamish 1976; Wright and Gjessing 1976; Beamish and Van Loon 1977), and these metals have been recorded in precipitation in these areas. Concentrations of alu-

minum and manganese, which are negligible in precipitation, also increase in many acidified

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678 HAINES

T^BLE 2.--Concentration (ixg/liter) of metals in lake waters of various acidities.

Locality (reference) a AI

Metal

Cu Cd Mn Ni Pb Zn

102 lakes, Ontario (average) (1) Blue Chalk Lake, Ontario (2) Lake Panache, Sudbury, Ontario (3) North Sweden (range) (4) Central Norway (range) (5) North Norway (range) (6)

13

(20-65

Nonacidified (pH 6.0-7.8) 2 (0.1 3 (3 (1 (1

8 40 3 9 6 28 6

0.05-0.23 (100 10-30

1-10 0-0.5 0-5 1-17

South-central Ontario, 14 lakes

(average) (2) 5.7 Nelson Lake, Ontario (7) 13 13

Intermediate (pH5.5-6.0)

Four lakes, Ontario (average) (1) 3 Clearwater Lake, Sudbury, Ontario (2) 453 97 Four lakes, Sudbury, Ontario (average) (8) 450 West coast Sweden (range) (4) 200-600 0.08-0.63 Southeast Norway (range) (5) 1-10 0-0.6 Lake Langtjern, Norway (average) (9) 218 6 0.21 South Norway (range) (10) 50-600 Adirondack lakes, New York (average) (11) 286 South Norway (range) (6) 40-600 Laxforsen, west Sweden (12) 288 1 0.2

49 3.6 12.6 18 10 16

Acidified (pH 4.1-5.3) 0.4 239 10 2 30

300 215 46

338 820 83 300400 1-5 30-122

1-10 3-35

2 15

45

190

23

3 28

a References: (1) Beamish 1976; (2) Dillon et al. 1979; (3) Yan et al. 1979; (4) Dickson 1975; (5) Wright and Gjessing 1976; (6) Wright et al. 1976; (7) Yan et al. 1977; (8) Adamski and Michalski 1975; (9) Henriksen and Wright 1977; (10) Wright and Snekvik 1978; (11) Schofield and Trojnar 1980; (12) Dickson 1980.

lakes and streams (Fisher et al. 1968; Dickson 1975, 1978, 1980; Beamish and Van Loon 1977; Almer et al. 1978; Cronan and Schofield 1979; Schofield and Trojnar 1980). The concentrations of these metals are inversely related to lake pH, and appear to reach these lakes because acidic precipitation leaches them h'om terrestrial soil and they are carried to the lakes by runoff and groundwater (Henriksen and Wright 1977; Cronan et al. 1978; Cronan and Schofield 1979). Iron, zinc, manganese, and aluminum are released from lake sediments

when an entire lake (Schindler2 et al. 1980) or enclosures within a lake (Schindler• et al. 1980) are intentionally acidified. Similarly, the addition of sulfuric acid to a New Hampshire stream caused an increase in the concentrations

of aluminum, calcium, magnesium, potassium, manganese, iron, and cadmium, but not of sodium, lead, nickel, chromium, or zinc (Hall and Likens 1980; Hall et al. 1980). Dickson (1980) found that in a group of Swedish lakes that received similar precipitation, cadmium, manganese, lead, and zinc were increased in lakes with low pH. Therefore, metals may increase

in acidified lakes from both direct input and dissolution from the watershed or sediments.

Galloway and Likens (1979) observed increased concentrations of gold, silver, cadmium, chromium, copper, lead, antimony, va- nadium, and zinc in the surface sediments of

two lakes in New York and one in New Hamp- shire. They concluded that these metals reached the lake, and thus the sediments, directly from the atmosphere, because elements that would be expected to be leached from the watershed (calcium, aluminum) did not follow this trend. However, this reasoning ignores that as pH declines, calcium and aluminum become more soluble and less is transported to the sediment.

The analysis of changes in metal concentration with depth in lake sediments presents a unique tool with which to examine chronolog- ical patterns of metal deposition. Oldfield et al. (1980) showed a dramatic acceleration in lead accumulation in the sediment of a south Bel-

gian pond during this century, steepening sharply from 1950. Taylor (1979) concluded that atmospheric input was responsible for en-

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richment of copper, lead, and zinc in surface sediments of Lake Windermere, England. Farmer et al. (1980) reached a similar conclu- sion concerning arsenic, cadmium, copper, chromium, lead, and zinc in sediments of Loch Lomond, Scotland, as did Norton et al. (1981) for lead and zinc in four lakes in New Hamp- shire and Maine, and Norton and Hess (1980) for these metals in 15 Norwegian lakes.

Seasonal Patterns

The input of contaminants to aquatic systems from polluted precipitation is not uniform, but varies seasonally and with individual storm events. The acid and metal content of precipitation from a particular storm depends on whether or not the storm has passed over industrial or urban areas (Elgmork et al. 1973; Dovland et al. 1976; Lunde et al. 1977). Indi- vidual storms have produced precipitation with a pH as low as 2.4 (Likens et al. 1979).

In Scandinavia, acid carried in precipitation is generally highest in autumn, and pH is depress?d in lakes and streams after fall storms (Gjessing et al. 1976). For example, the pH of Langtjern Lake, Norway, declined from 5.2 in August to 4.7 in October (Henriksen and Wright 1977). In North America, acid input from precipitation is generally highest in summer and lowest in winter (Hornbeck et al. 1976); however, the pollutants contained in snow and stored in the snowpack are released during spring snowmelt, sometimes resulting in severe pH depressions in surface waters. After spring snowmelt, pH declined from 5.2 to as low as 4.0 in the Tovdal River, Norway (Lei- vestad and Muniz 1976), and from 5.6 to 3.8 in Honnedaga Lake, New York (Schofield 1973). In Ontario, 36-77% of the total annual export of hydrogen ion from watersheds studied occurred in April, resulting in a 2- to 13-fold increase in hydrogen ion concentration in streams (Jeffries et al. 1979).

Other pollutants contained in snow, such as heavy metals, also are released during melting and consequently increase in surface waters (Hultberg 1977). In New York, the acid released from melting snow reduced stream pH from about 5.5 to 4.5, and increased aluminum

concentration from 0.2-0.4 mg/liter to as much as 1.0 mg/liter (Schofield and Trojnar 1980). As there was little aluminum in the snow, the acid must have dissolved aluminum from unfrozen

terrestrial soil or aquatic sediment. Seasonal patterns of low pH and high concentrations of free metals were observed in several lakes near

Sudbury, Ontario as a result of influx of meltwater (Scheider et al. 1976).

In Norway, the first 30% of meltwater contains 50-80% of the total amounts of acids, metals, and all other substances measured in

the snowpack (Gjessing et al. 1976; Hultberg 1977; Johannessen and Henriksen 1978; Skart- veit and Gjessing 1979). This phenomenon has been attributed to ion separation during melting (Hultberg 1977) or to a freeze-concentration process (Johannessen and Henriksen 1978). A similar pattern was observed in snow from northern Minnesota (Glass 1980; Siegel 1981). Johannessen et al. (1980) described three phases of stream-chemistry changes related to snowmelt: input of groundwater; peak ion concentration from snow; and dilution from remaining snow. However, in Ontario the release of acid from melting snow was directly proportional to the percent melted (jeffries et al. 1979). In both situations, however, pH declined and metal concentrations increased in

surface waters as a result of snowmelt.

Changes in Non-Fish Aquatic Biota Bacteria and Fungi

Early investigators reported an accelerated accumulation of coarse organic detritus in acid Swedish lakes, and the occurrence of dense mats of fungus hyphae on the sediments of some lakes (Grahn et al. 1974). It was hypoth- esized that decomposition in these acid lakes was done primarily by fungi, rather than by bacteria, and that the detrital accumulation was the result of a reduced rate of decomposition. Hendrey et al. (1976) and Leivestad et al. (1976), who reviewed studies on the effects of acidification on decomposers in Norway, reported that the rate of decomposition of birch leaves was reduced in both field and laboratory studies when water pH was below 6.0. In a res- pirometer study in which homogenized leaf lit- ter was used as a substrate, oxygen consumption was halved when pH was reduced from 7.0 to 5.2. When the experiments were repeated with glucose and glutamic acid substrates, decreased pH led to a shift in dominance of organisms from bacteria to fungi, and a decrease in oxygen consumption. Similar results were

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680 HAINES

obtained by Traaen (1980). Minshall and Min- shall (1978) found that decomposition of detritus was 5-20 times faster in streams with pH 5.6-7.1 than in streams with pH 4.6-5.4

Recently, Hendrey and Vertucci (1980) have reported that earlier reports of development of fungal mats in acidic lakes were incorrect, and that mats actually are composed of filamentous algae. Although the fungal mat may not occur, fungi do increase in acidified lakes. Hall and Likens (1980) and Hall et al. (1980) found that hyphomycete fungal density decreased in an experimentally acidified stream, but basidio- mycete fungal density increased.

In a laboratory study with sediment-water microcosms from Swedish lakes, Andersson et al. (1978) found that microcosms from lakes with pH about 5 had reduced sediment oxygen consumption and carbon dioxide production, and increased glucose turnover time, as compared to lakes with pH about 6.8, indicating reduced microbial activity at reduced pH. These findings are contradicted by those of Gahnstrom et al. (1980) who found no differences in glucose turnover time or oxygen consumption in microcosms with profundal sediments from lakes with pH 4.5-5.2 as compared' to lakes with pH 6.5-6.8.

Schindler2 et al. (1980) intentionally acidified a lake, reducing the pH from 6.7-7.0 to 5.7- 5.9, but found no evidence that the decomposition of organic detritus was reduced after acidification. The addition of sulfuric acid apparently caused an increase in sulfate-reducing bacteria. Muller (1980) acidified enclosures in the same lake and found no effect of acid on

bacteria numbers. Scheider et al. (1975) reported that the open-water microbial populations in acidified lakes near Sudbury, Ontario were markedly different from those in similar, nonacidified lakes, but the sediment microbial

populations were not. Among open-water mi- crobes, ammonia-oxidizing bacteria were not found in acidified lakes (indicating reduced nitrogen cycling) and aerobic heterotrophs, sulfate reducers, nonacidophilic sulfur oxidizers, and coliforms were scarcer than in nonacid

lakes. Acidophilic sulfur-oxidizing bacteria, on the other hand, were more abundant in acidified lakes, as might be expected. Yeast and mold populations were similar in all lakes. But Traaen (1980) found no differences in planktonic bacteria in acid and nonacid lakes, and

concluded that concentration of dissolved or-

ganic matter was the major factor affecting planktonic bacteria biomass.

When the acidified lakes were chemically neutralized, Scheider et al. (1975) found that open-water microbial populations quickly responded and soon resembled those in the nonacid lakes. Hendrey et al. (1976) reported that lime treatment of acid Swedish lakes caused a

rapid decomposition of organic detritus and reductions in fungal mats (which were probably algae mats). Gahnstrom et al. (1980) found that addition of lime to one acidic lake resulted in

increased rates of glucose turnover and oxygen consumption in the sediment, even though the microcosm study before lime was added indicated that these rates were similar to those in

nonacid lakes.

Acidification of lakes and streams to pH 5 or less appears to reduce decomposition because of the elimination of some species of bacteria, which may be replaced by other species of bacteria or fungi. The addition of lime to raise lake pH increases decomposition. There are many discrepancies among the above studies that may be the result of sample-handling techniques that change microbial populations.

Primary Producers

Acidification affects communities of planktonic and benthic algae in lakes and streams. The number of species of phytoplankton in acidic lakes decreases as pH declines (Almer et al. 1974, 1978; Hendrey et al. 1976; Kwiat- kowski and Roff 1976; Leivestad et al. 1976; Yan and Stokes 1978; Yan 1979). Similar effects were observed for periphyton in artifi- cially' acidified enclosures (Muller 1980). Dia- tom samples collected from a series of acidified lakes in Norway showed that the proportion of acidophilic species increased between 1949 and the mid 1970s (Leivestad et al. 1976). In Dutch moorland pools, diatom species diversity declined from 1920 to 1978 in acidified pools, but not in nonacidified pools (Van Dam et al. 1980), and the number of desmid spedes also declined (Coesel et al. 1978).

Although the number of species is reduced, phytoplankton biomass and production may be similar in acid and nonacid lakes with similar

phosphorus levels (Hendrey et al. 1976; Almer et al. 1978; Yan and Stokes 1978; Dillon et al. 1979; Yan 1979; Raddum et al. 1980). Phos-

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phorus levels frequently decline at low pH, but they may increase at pH 4.0 or less (Almer et al. 1978). The addition of phosphorus to acidic lakes in Ontario resulted in an increase in phytoplankton biomass (Dillon et al. 1979), but De- Costa and Preston (1980) could not stimulate phytoplankton growth at pH below 5.5 in a lake acidified by mine drainage, unless the water was buffered. Schindler (1980) found that phytoplankton biomass and production did not decline when a lake was experimentally acidified from pH 6.6 to 5.6.

Phytoplankton species composition also changes as pH declines. In Swedish lakes with pH above 6, phytoplankton populations contained all groups in more or less equal numbers. As pH decreased, the Chrysophyta, Cy- anophyta, and Chlorophyta were reduced; at pH about 4 Pyrrhophyta became dominant (Almer et al. 1974, 1978). The same pattern was found in Ontario lakes (Yah and Stokes 1976, 1978; Yah 1979); even some of the same species were dominant in acidic lakes in both countries. In contrast, Kwiatkowski and Roff (1976) found that Cyanophyta became dominant in six Ontario lakes as pH declined, a trend opposite to that observed by others. Also, Schindlers et al. (1980) found no changes in phytoplankton species composition in a lake after intentional acidification; however the pH was reduced only from 6.6 to 5.6.

Similar to the situation for planktonic algae, the standing crop of periphytic algae has increased in some acid lakes and streams even

though the number of species present is reduced and the species composition is changed (Hendrey et al. 1976; Leivestad et al. 1976). Ahner et al. (1978) and Nilssen (1980) report that the filiform alga Mougeotia sp. (Conjuga- tae) frequently is fbund in large masses in acidic lakes and streams. This alga also became abundant in an Ontario lake that was experimentally acidified, first appearing in samples at pH 5.6 (Schindler 1980). Similar results have been produced in streams and in lake enclosures after

artificial acidification (Hendrey 1976; Hall and Likens 1980; Hall et al. 1980; Muller 1980). In these experiments, total primary production either did not change at low pH, or declined. The increase in algal standing crop at low pH may, therefore, be explained by reduced heterotrophic activity--both microbial and invertebrate.

Aquatic macrophytes have shown changes related to acidification. In Swedish lakes,

macrophyte communities once dominated by Lobelia sp. were later dominated by Sphagnum sp. (Grahn et al. 1974; Grahn 1976) or Juncus bulbosus (Nilssen 1980). A simlar dominance of sphagnum was evident in an acidic lake in New York (Hendrey and Vertucci 1980). In a laboratory study, the growth and production of Lobelia sp. was reduced at pH 4 (Leivestad et al. 1976; Hendrey et al. 1976). Grahn et al. (1974) and Hendrey and Vertucci (1980) believe that the increase of Sphagnum sp. may ac- celerate the acidification process because Sphagnum has a high ion-exchange capacity in the cell walls.

Invertebrates

Both planktonic and benthic invertebrate communities are affected by acidification of lakes and streams. In Ontario and Sweden, zooplankton biomass was lower in acidic than in similar, nonacidic lakes (Roff and Kwiatkowski 1977; Ahner et al. 1978; Dillon et al. 1979; Yan and Strus 1980). The number of zooplankton species present in a lake decreased as pH decreased in several areas: Sweden (Almer et al. 1974, 1978); Norway (Hendrey and Wright 1976; Raddum 1980); and Ontario (Sprules 1975a, 1975b; Roffand Kwiatkowski 1977; Yan and Strus 1980). In Ontario, lakes with pH above 5 contained zooplankton communities of 9-16 species, with three or four dominant; lakes with pH below 5 contained 1-7 species, with only one or two dominant. Compared with morphometrically similar nonacidified lakes in northwest Ontario, acidified lakes near Sud- bury had unusually simple zooplankton communities (Sprules 1975a).

The acidic lakes in both Scandinavia and

Canada are characterized by a dominance of bosminids and a scarcity of daphnids and rotifers (Ahner et al. 1974, 1978; Scheider et al. 1975; Sprules 1975a; DeCosta and Preston 1980; Raddum 1980; Yah and Strus 1980). Acidic lakes frequently are dominated by species of Bosmina or Diaptomus. Certain species of rotifers are acid-tolerant. For example, Ker- atella serrulata occurs in nearly every acidic Nor- wegian lake (Raddum 1980).

Low pH affects survival, reproduction, and distribution of some zooplankton species. Davis and Ozburn (1969) found that exposure to pH

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682

5 reduced survival of Daphnia pulex, and there was no survival at pH 4.3 or less. Reproduction did not occur below pH 7. Daphnia magna does not reproduce at pH 5.5 or below (Parent and Cheetham 1980). Groterud (1972) reported that zooplankton avoided low-pH surface water in a lake. Lowndes (1952) gives the pH ranges at which many species of entomostracans occur.

Benthic invertebrate populations are impoverished in acidified environments (Almer et al. 1978). Okland and Okland (1980) concluded that the number of species of macroinverte- brates declined in Norwegian lakes as pH declined below 6.0. Mossberg and Nyberg (1979) found that the number of species declined with pH in lakes with pH of 4.2 to 5.0. A lake in Sweden with pH 3.9-4.6 had a low standing crop of zoobenthos and few species (Wieder- holm and Eriksson 1977). Reducing the pH of a stream from 5.4 or higher to 4.0 resulted in reduced standing crop and diversity of stream zoobenthos (Hall and Likens 1980; Hall et al. 1980).

Mollusks are highly sensitive to acidification, as would be expected due to the high CaCO3 requirement of this group for shell formation. Snails were not found in Norwegian lakes with pH at or below 5.2, and were rare or reduced at pH 5.2-6.6 (Okland 1969; J. Okland 1980a, 1980b; Raddum 1980). The snail Ancylus sp. was not found below pH 5.7 in the River Dud- don, England (Sutcliffe and Carrick 1973). The number of species of fingernail clams (Sphae- riidae) declined at low pH in Norwegian lakes (Okland 1971). Of 20 species, only 6 are found in lakes with pH less than 5 (K. Okland 1980). Mollusks are not found in Ontario lakes with

pH at or below 5 (Scheider et al. 1975; Scheider and Dillon 1976; Roff and Kwiatkowski 1977).

The crustaceans Gammarus lacustris and Lep- idurus arcticus are widespread in Norway and are important fish-food organisms where they occur. They are not found in lakes with pH less than 6.0 (Leivestad et al. 1976; K. Okland 1980). Gammarus lacustris was not found below pH 5.7 in the River Duddon, England (Sutcliffe and Carrick 1973). Borgstrom and Hendrey (1976) found that exposure to pH at or below 5 caused 80-100% mortality in Gammarus lacustris and delayed or halted molting in Lepi- durus arcticus. Minshall and Minshall (1978) found that the mortality rate of Gammarus lacustris was twice as high in water of pH 5.0-6.0

than at pH 6.3-7.0. Costa (1967) found that Gammarus pulex avoid water of pH 6.2 and lower.

In Sweden, the crayfish Astacus astacus is common in lakes with pH greater than 6.4 but is rare in lakes with pH less than 6.0. This species is commonly harvested for use as food (Almer et al. 1978). Cambarus longulus can survive pH 3 for 10 days, but not pH 2; however Cambarus latimanus is not found at pH 5.6 or below (Hobbs and Hall 1974). In crayfish exposed to low pH, calcium uptake was inhibited at pH below 5.75, and the progression of molt stages and calcification of the exoskeleton were re-

duced at pH 5.0 (Malley 1980). Some groups of aquatic insects are reduced

at low pH while others flourish. Many species of Ephemeroptera and Plecoptera disappear as pH declines. In Norway, 10 of 22 species in these two groups were highly correlated with lake pH. At pH 6.5 or above there were 8-12 species present. This declined to 1-2 at pH 4.0- 4.5 (Hendrey and Wright 1976). Ephemerop- tera were reduced in Norwegian lakes with pH 4.2-5.2 (Nilssen 1980), and were absent at pH below 5.7 in the River Duddon, England (Sut- cliffe and Carrick 1973). In the River Duddon, eggs of Baetis sp. were found only where pH was 6 or higher. Scheider et al. (1975) and Scheider and Dillon (1976) report that Ephem- eroptera were absent from acidic lakes in On- tario.

In a laboratory study, Bell (1971) found that Ephemeroptera were intolerant of low pH, and Plecoptera were moderately tolerant. The acidification of a stream in New Ha•npshire from pH 5.4 or greater to 4.0 caused reduced emergence of Ephemeroptera and some Plecoptera (Hall and Likens 1980; Hall et al. 1980). In an earlier study in the same stream, Fiance (1978) reported that reducing the pH from 5.0-6.2 to 4.0-5.4 did not affect emergence of adult Ephemerellafuneralis, but did cause a decrease in growth and recruitment.

Various species in other groups of aquatic insects may also be sensitive to reduced pH. In the River Duddon, England, the trichopterans Wormaldia sp. and Hydropsyche sp. were not found below pH 5.7. Raddum (1980) reported that chironomids were reduced at low pH in Norwegian lakes. Hall and Likens (1980) reported that the dipterans Orthocladinae and Tanypodinae were reduced when strea•n pH

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ACIDIC PRECIPITATION; DEGRADATION OF AQUATIC ECOSYSTEMS 683

was reduced. In Ontario, Chaoborus sp. were absent from acidic lakes (Scheider et al. 1975; Scheider and Dillon 1976). Bell (1970) found that the life cycle of the midge Tanytarsus dissimilis could not be completed below pH 5.5.

Other groups of aquatic insects may become very abundant at low pH. Raddum (1980) found that Coleoptera, Hemiptera (Corixidae), and Megaloptera were more abundant in lakes with pH below 4.8 than in circumneutral lakes. Nilssen (1980) found Hemiptera and Coleop- tera at increased abundance in Norwegian lakes as pH declined. Mossberg and Nyberg (1979) found Odonata, Heteroptera, and the dipterans Chaoborus sp. and Chironomus sp. to be very abundant in lakes with pH 4.2-5.0. This response of dipterans is the opposite of that found in Norway by Raddum (1980) and in North America (Scheider and Dillon 1976; Hall and Likens 1980). Bell (1971) found Trichop- tera to be tolerant of reduced pH and Odonata to be moderately tolerant.

Few reports exist for other groups of aquatic invertebrates, although these groups may be equally as important as the better-studied groups in ecosystem functioning. Among the aquatic Annelida, Oligochaeta are reported to be rare in the deeper areas of acidic Norwegian lakes (Leivestad et al. 1976; Raddum 1980), and Hirudinea were not found below pH 5.5 (Nils- sen 1980; Raddum 1980). In a review of the Hirudinea, Sawyer (1974) found few reports of leeches occurring below pH 6. Certain groups of sponges require CaCO.• for spicule formation. Jewell (1939) found that calcium was an important factor in the distribution of species of sponge in Wisconsin. Racek (1969) found that the soft, acidic waters in eastern Australia and the hard, alkaline waters in western Aus-

tralia harbored very different species of sponge. It is probable that acidic precipitation may affect sponge distribution and abundance, but no evidence of this currently exists.

The effects of acidification on invertebrates

may result from direct physiological effects of acids or toxic metals on various organisms, through alteration in food supply, or through changes in predator-prey relations. Bell (1971) found that pH was directly lethal to some benthic invertebrates. Emergence was the most critical period. The lowest pH at which 50% of the adults emerged was 4.0 for the most tolerant species (Trichoptera) and 5.9 for the least

tolerant (Ephemeroptera). Similar laboratory results were obtained by Bell and Nebeker (1969) and Butler et al. (1973). Exposure to low pH causes a reduction in sodium uptake and an increase in sodium loss from sensitive inver-

tebrates, including mollusks, crustaceans, and insects (Potts 1979; Potts and Fryer 1979). These effects are the same as found in sensitive

fish (see below). Conversely, exposure to pH as low as 3.0 did not affect serum sodium or chlo-

ride levels in an acid tolerant invertebrate Co-

rixa sp. (Vangenechten and Vanderborght 1980).

Raddum (1980) believes that toxic metals may be important in determining the effect of acidification on invertebrates. He observed that

sediments, the habitat of many invertebrates, had a higher pH than the overlying lake water, and that oligochaete abundances were increased in acidic lakes in shallow water, but reduced in deep water. He concluded that toxic metals were responsible for this response, but presented no data to verify this. Invertebrates are susceptible to heavy metal toxicity, and concentrations of some metals (for example, lead, zinc) are increased in surface sediments of acidic lakes (Norton and Hess 1980; Norton et al. 1981). This area needs more investigation.

The possibility that alteration of food supply in acidic waters may affect aquatic invertebrates is suggested by functional group analysis of the invertebrate species affected by acidification. Sutcliffe and Carrick (1973) observed that most of the species that were reduced in acid streams were herbivores, or scrapers. Friberg et al. (1980) found that scraper species were three times as abundant in streams of pH 6.5-7.3 than at pH 4.3-5.9, whereas shredder species were twice as abundant in the low pH streams. This suggests that a reduction in periphyton available for use as food may be responsible for the reductions in these invertebrates. Sutcliffe

and Carrick (1973) reported an observed reduction in algae in acid sections of the River Duddon, but this was not quantified. However, Minshall and Minshall (1978) found that the toxicity of acid water to Baetis sp. and Gammarus sp. was the same whether the organisms were fed or not. These discrepancies show the dif- ficulty of determining cause and effect in complex ecosystems when only a few variables are measured.

Eriksson et al. (1979) and Henrikson et al.

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684 I•^•Es

(1980) believe that the reported effects of acidification on algae and invertebrates are the re- suit of predator-prey relations resulting from the disappearance of fish, and not of toxic effects of acids and metals on these organisms. They produced similar effects, such as reduced number of species and altered species dominance, by experimentally removing fish from a nonacidic lake. Henrikson and Oscarson

(1978) found that fish predation reduced abundance of the waterbug Glaenocorisa p. propinqua, and this insect has recently colonized lakes in south Sweden, where fish have been eliminated

by acidification. Raddum (1980) reported that average size of invertebrates increased in the absence of fish. However, these observations

cannot explain the observed responses of algae and zooplankton in Ontario lakes when fish still were absent after chemical neutralization

(Scheider et al. 1975; Dillon et al. 1979). Roff and Kwiatkowski (1977) reported that zooplankton size distribution was uniform in six Ontario lakes even though fish were absent from some.

Amphibians

Amphibians also are affected by acidic precipitation. Glass and Loucks (1980) cited a study that documented the decline of the frog Rana temporaria and the toad Bufo bufo from a Swed- ish lake where the pH had declined to 4.0-4.5, and from which all fish had disappeared. Pough (1976) cited studies indicating that a decline in British frog populations is related to acidification of breeding sites. Gosner and Black (1957) found reduced numbers of frog species in the acidic waters of the New Jersey pine barrens, and found that only 4 of 11 species could reproduce at pH 4.1 or lower.

Pough and Wilson (1976), who determined the effect of pH on hatching success in two species of mole salamander, observed that one of these, Ambystoma maculatum, did not tolerate pH below 6. Temporary ponds in New York used by this species for reproduction had a mean pH of 4.5 (range 3.5-7.0), and a field study showed high mortality of this species in ponds of low pH (Pough 1976; Pough and Wil- son 1976). However, Cook (1978) found low mortality rates and no correlation between mortality of this salamander and pH of breeding ponds over a pH range of 4.0 to 6.0 in

Massachusetts. Huckabee et al. (1975) placed larval shovelnose salamanders Leurognathus marmoratus in containers in a creek above and

below a pyritic road fill. Above the fill creek, pH was 6.9-7.2 and aluminum concentration was less than 0.01 mg/liter; below the fill, the pH was 4.5•1•.9 and aluminum concentration was 1.0 mg/liter. No salamanders died above the fill but most died below it. Inasmuch as sim-

ilar pH and aluminum concentrations occur in lakes acidified by precipitation, salamanders may be adversely affected in such waters.

The mechanisms by which amphibians are affected by acid stress may be similar to those for fish (see below). Exposure to acid decreased sodium influx in isolated frog skin, and thereby reduced active sodium transport (Po Fromm, Michigan State University, personal communication). However, there was no significant change in osmotic permeability of intact frogs Rana pipiens exposed to low pH.

Changes in Fish Populations

Effects of acidic precipitation on fish include mortality, reproductive failure, reduced growth rate, skeletal deformities, and increased uptake of heavy metals. The earliest recorded impact of acidic precipitation on aquatic systems was a decline in Atlantic salmon • in a few southern

Norway rivers in the 1920s, which was correlated with low pH in these rivers (Jensen and Snekvik 1972; Wright et al. 1976). These declines have continued, and the Atlantic salmon catch has declined to virtually zero in seven southern rivers that are subject to acid precipitation and now have pHs generally between 4.5 and 5.5. No such trend was evident in 68

northern rivers that are not being acidified (jensen and Snekvik 1972; Leivestad et al. 1976; Wright et al. 1976; Overrein et al. 1980; Muniz 1981). In hatchery experiments, neutralization of water from one of the affected

rivers with sodium hydroxide or limestone re- suited in normal hatching and survival of At-

• Scientific names of most fish species mentioned in this review are given in Table 3. Others: channel cat- fish Ictalurus punctatus; common carp Cyprinus carpio; European cisco Coregonus albula; fathead minnow Pimephales promelas; flagfish Jordanella floridae; goldfish Carass'ius auratus.

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lantic salmon, indicating that low pH was the cause of the mortality. Efforts to reestablish Atlantic sahnon runs by stocking hatchery fish have failed (Leivestad et al. 1976).

A survey of more than 2,000 lakes in southern Norway, begun in 1971, revealed that about one-third had lost their fish populations (pri~ rnarily brown trout) since the 1940s (Wright and Snekvik 1978). In a sample of 700 lakes for which both fish population and water chemistry data were compiled, fish population status was related to pH. Populations were sparse or absent at pH 5.5 or less. In about 80-85% of the lakes, pH was below 5. Fish populations have been affected in lakes and streams in an area

of 33,000 krn 2 (Overrein et al. 1980; Sevaldrud et al. 1980).

In 68 lakes in the LaCloche Mountains, On-

tario surveyed by Harvey (1975), the number of species of fish present decreased as lake pH decreased. The disappearance of several fish species from various lakes in this region has been reported (Beamish and Harvey 1972; Bearnish 1974b, 1976; Bearnish et al. 1975). Harvey (1980) reported that fish have been lost from about 200 Ontario lakes. In southwestern

Nova Scotia, pH is less than 5 in nine of 16 rivers and less than 6 in five. Historical water-

chemistry data show that the pH of at least six of these rivers has declined since 1955, and that sulfate has increased (Altshuller and McBean 1979). Seven rivers in western Nova Scotia are now believed to be unsuitable for Atlantic sahn-

on reproduction, and another eight are threat- ened (Farmer et al. 1980). Angler harvest records for two of the unsuitable rivers show that

catch was relatively constant from the 1930s to the 1950s, then declined to virtually zero. In contrast, angling catches in four other rivers currently having good water quality have re- rnained relatively constant (Farmer et al. 1981).

In 217 lakes above 610 rn elevation in the

Adirondack Mountains of New York, pH was less than 5 in 111, and 100 of these had no fish. When 40 of these lakes were surveyed in 1929- 1937, only 4% had pH below 5 and no fish (Schofield 1976). In 849 Adirondack lakes surveyed in 1975-1979, pH was less than 5 in 212 and between 5 and 6 in 256 (Pfeiffer and Festa 1980). At least 113 lakes that once supported brook trout were devoid of fish. A number of

lakes that formerly contained smallmouth bass

had lost this species as the result of acidification.

Mortali• Acute mortalities of fish observed in acidified

waters have occurred primarily in streams, and have been associated with an event that caused

rapid change in water pH, such as snowmelt or heavy autumn rains. Mortalities of Atlantic sahnon and brown trout have been recorded in

Norway (Jensen and Snekvik 1972; Leivestad and Muniz 1976; Leivestad et al. 1976; Wright et al. 1976). The pH measured during these die-offs ranged from 3.9 to 4.6. Atlantic sahnon fry died in hatchery pools fed with water from the Mersey River, Nova Scotia, the pH of which was 5.0 (Farmer et al. 1980). Brook trout died during spring snowmelt in a New York laboratory that was rearing fish in water piped from a nearby stream (Schofield and Trojnar 1980). The lowest pH was about 5.2, and aluminum concentration reached 1 rng/liter. Large fish kills during summer in two Swedish lakes were attributed to high aluminum concentrations from acidic spring runoff and sudden pH elevation caused by primary production (Grahn 1980).

A sudden decrease in pH or increase in aluminum associated with some climatic event may cause fish mortality, because rapid changes in toxic substances can cause death at much lower

concentrations than chronic exposure to gradually increased concentrations. Death of fish also occurs at higher pH in water with very low ionic content than in water with moderate or

high ionic content (Grande and Andersen 1979; Fromm 1980; Rosseland et al. 1980). Ion- ic content of surface waters is also generally lowest during spring snowmelt or heavy precipitation.

Death of fish at low pH has been attributed to failure of ion regulation or asphyxiation, or to elevated metal concentrations associated with

low pH. Exposure to 1ow-pH water causes ede- ma between outer gill lameliar cells and re- rnaining tissue, erosion of larnellae, and swell- ing of filaments (McKenna and Duerr 1976). Pennsylvania State University (1971) research- ers concluded that the primary mode of acid toxicity in fish is gill damage, which impairs respiratory, excretory, and liver functions. Liver impairment reduced tolerance of fish to other

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toxicants. Daye (1981) believes that the death of fish embryos is the result of corrosion of epi- dermal cells by acid, which interferes with respiration and osmoregulation.

Brown trout collected from the Tovdal River, Norway, during an acid-caused fish kill had lower plasma sodium and chloride concentrations than did fish from unaffected sections of

the river. The reduced concentrations were

comparable to those found in fish stressed by low pH in a tank experiment (Leivestad and Muniz 1976). When brown trout were trans- ferred from the Tovdal River to a more acidic

tributary, plasma sodium and chloride concentrations declined rapidly, and mortality began after 11 days (Muniz and Leivestad 1980a). A decrease from pH 7 to pH 4 caused a threefold increase in sodium loss from brown trout--suf-

ficient to cause death in 24-48 hours (Mc- Williams and Potts 1978; Potts 1979). As pH decreased below 6, sodium uptake decreased and sodium loss increased (McWilliams et al. 1980). Exposure to increased hydrogen ion concentration increases gill membrane permeability. Hydrogen ions from the environment flood in and sodium and other ions from the

blood flood out (Packer and Dunson 1970; Fromm 1980). This results from a failure of active uptake of sodium by the gill chloride cells, and movement of ions in relation to concentration gradients (Spry et al. 1981). Gill membrane permeability to monovalent ions is mediated by calcium (and possibly other diva- lent cations) whose presence in the water reduces permeability. Thus, as the calcium concentration in water increases, the loss of ions is

reduced and the level at which pH becomes lethal decreases (Evans 1975; Oduleye 1975; McWilliams and Potts 1978; McDonald et al. 1980; Brown 1981). Calcium uptake in fish also occurs in the chloride cells (Payan et al. 1981). Despite the influx of hydrogen ion, blood acidosis does not occur in fish when the water is

low in calcium. As only a small amount of ionic hydrogen can be excreted by the kidney, the remaining ionic hydrogen apparently accumu- lates and is buffered intracellularly (Spry et al. 1981).

Low pH may interfere with respiration through several mechanisms. Elevated hydrogen ion concentration may cause excessive se- cretion of mucus from the gills, thereby reduc-

ing the rate of oxygen diffusion across the gill surface (Plonka and Neff 1969; Vaala et al. 1969; Daye and Garside 1976; Dively et al. 1977; Ultsch and Gros 1979). At extremely low pH, an increased influx of hydrogen ions reduces blood pH which, in turn, reduces the oxygen-carrying capacity of hemoglobin. This effect is most pronounced in water with high calcium concentration (Spry et al. 1981). Packer (1979) reported reduced oxygen consumption in brook trout exposed to acutely lethal pH. The reduced consumption was caused by decreased oxygen transfer and reduced blood oxygen capacity.

In response to chronic acid exposure, the hematocrit index, hemoglobin content of blood, and hemopoietic activity of fish are all increased, presumably to maintain oxygen-carrying capacity (Neville 1979a, 1979b; Fromm 1980). Vaala and Mitchell (1970) found reduced blood oxygen tension and compensatory erythrocytic changes in brook trout exposed to pH 34. These compensatory mechanisms may enable fish to breathe normally after they are acdimated to low pH. Exposure to low pH did not change oxygen consumption by brown trout (Carrick 1981), bluegills, or goldfish (Ultsch 1978). However, exposure to extremely low pH may override compensatory mechanisms. Thus Ultsch (1978) found decreased oxygen consumption at pH 3.5 for channel cat- fish, and Packer and Dunson (1972) reported decreased oxygen consumption prior to death for brook trout at pH less than 3.5.

The most likely explanation of the physiological effect of hydrogen ion on fish is that at moderately low pH (4 to 5), failure of ion regulation is the primary response. At very low pH (•<3.5), respiratory failure occurs. Ion-regulation failure is probably the most widespread response in acidified waters, which rarely have a pH less than 4. The physiological changes that have been ascribed to respiratory acclimation (for example, increased hematocrit, hemoglobin, and hemopoiesis) may in fact be ion-reg- ulatory adjustments to cope with increased in- ternal hydrogen ion (Spry et al. 1981).

Acidification of surface waters is accom-

panied by increases in concentrations of some metals. Aluminum concentration appears to be very important in determining the effect of acidification on fish. Cronan and Schofield

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(1979), Baker and Schofield (1980), and Scho- field and Trojnar (1980) showed that mortality of brook trout in New York was caused by aluminum and pH in combination, rather than by either factor singly. Similar results were reported by Grahn (1980) for European cisco, by Herrmann and Baron (1980) for brook trout, and by Leivestad et al. (1980) and Muniz and Leivestad (1980b) for brown trout.

The toxicity of aluminum varies with pH and the presence of complexing agents. For aluminum in the hydroxyl form, toxicity is highest at pH 5, and declines at both higher and lower pHs. Aluminum complexed with organic matter is not toxic to organisms (Baker and Scho- field 1980; Driscoll et al. 1980). At pH 5, aluminum concentrations of 0.2 mg/liter or greater caused gill hyperplasia and mucus se- cretion in b•*ook trout (Schofield and Trojnar 1980). Exposure to aluminum concentration of 0.15 mg/liter at pH 5 caused loss of blood ions and clogging of gills with mucus in brown trout (Leivestad et al. 1980; Muniz and Leivestad 1980b). Fish exposed to acidified tap water did not show these effects, but those exposed to water from acidified streams did. Rosseland

(1980) found that brown trout exposed to tap water acidified to pH 4.5 showed no increase in standard metabolism or respiratory rate, but those exposed to pH 4.5 stream water did. Ad- dition of aluminum to the acidified tap water produced the same responses. Baker and Scho- field (1980) found that white suckers were more sensitive to aluminum than brook trout, and that in both species sensitivity to pH decreased with age but sensitivity to aluminum increased. Thus the factor responsible for death may change with age.

Yah et al. (1979) reported mortality of rainbow trout stocked in an acidified lake that had

been chemically neutralized. Fish had been eliminated from the lake by acidification. The mortality was attributed to copper, or copper and zinc in combination. Waiwood (1980) reported increased hematocrit in rainbow trout exposed to 42-67 txg copper/liter at pH 6. Cop- per concentrations as low as 9.5 txg/liter were reported to be toxic to brook trout embryos and juveniles in laboratory exposures (McKim and Benoit 1971). Copper concentrations up to 450 /xg/liter have been measured in lakes near Sud- bury, Ontario, which were devoid of fish

(Adamski and Michalski 1975). The toxicity of many metals increases as pH and calcium decline (Waiwood and Beamish 1978; Chrost and Pinko 1980; McFarlane and Franzin 1980).

In addition to acute mortality of fish, gradual mortality of adults during long periods of exposure to reduced pH has been reported. Beamish (1974a) observed selective mortality of large white suckers in Lumsden Lake, Ontario as lake pH declined over several years. Franzin and McFarlane (1980) reported similar mortalities. Rosseland et al. (1980) found that in some cases, brown trout and European perch populations in acidic lakes show a lack of older fish.

The fish seemed to die after spawning, possibly as a result of additional stress related to spawning.

Reproduction Although fish can die from acidification,

more commonly they fail to reproduce. Re- cruitment ceases and a species eventually disappears from the acidified lake or stream. Var- ious species of fish have disappeared from natural populations at different apparent pHs (Table 3). The apparent pH at which the species disappears is higher than the pH known to be acutely toxic to adult fish. In short-term laboratory studies of the effect of hydrogen ion on fish (Table 4), adult fish were more resistant than any other life stage to reduced pH. Thus, it is likely that the loss of fish populations from acidified lakes is, in most cases, a result of reproductive failure rather than of mortality of adult fish, although exceptions may occur.

Several mechanisms are involved in causing apparent failure to reproduce. In toxicity tests with acids and metals, early life-history stages generally are more sensitive than the adults of the same species (McKim 1977). For example, adult brook trout can survive at pH 3.5-4.5, but embryo mortality is increased at pH 4.5- 6.5, and fry mortality at pH 4.4-6.1 (Table 4). Copper concentrations above 23 /xg/liter were toxic to adult brook trout, but concentrations as low as 9.5 txg/liter were toxic to embryos and juveniles (McKim and Benoit 1971). Huckabee and Griffith (1974) found that mercury concentrations of 3/xg/liter or greater reduced the hatch of eggs of common carp, and that the reduction was exacerbated by the presence of selenium. Horning and Neiheisel (1979) re-

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T^•LE 3.--Species of fish that ceased reproducing, declined, or disappeared from natural populations as a result of acid•tication from acidic precipitation, and the apparent pH at which this disappearance occurred.

Apparent pH at which population ceased reproduction, declined, or disappeared

Family and species (reference) a

Salmonidae

Lake trout Salvelinu• namaycush Brook trout Salvelinusfontinali• Aurora trout Salvelinus fontinalis timagamiensis Arctic char Salvelinus alpin• Rainbow trout Salmo gairdneri Brown trout Salmo trutta

Atlantic salmon Salmo salar

Lake herring Coregonus artedii Lake whitefish Coregon• clupeaformis

Esocidae

Northern pike Esox lucius'

Cyprinidae Golden shiner Notemigonus crysoleucc• Common shiner Notropis cornutus Lake chub Couesi• plumbeus Bluntnose minnow Pimephales notat• Roach Rutilus rutilus

Catostomidae

White sucker Catostom• commersoni

Ictaluridae

Brown bullhead lctalurus nebulosu•

Percopsidae Trout-perch Pefcopsis omiscomaycu•'

Gadidae

Burbot Lota Iota

Centrarchidae

Smallmouth bass Micropterus dolomieui Largemouth bass Micropter• salmoides Rock bass Ambloplites rupestris Pumpkinseed Lepomis gibbos• Bluegill Lepomis macrochirus

Percidae

Johnny darter Etheostoma nigrum Iowa darter Etheostoma exile

Walleye Stizostedion v. vitreum Yellow perch Percaflavescens European perch Percafluviatilg•

5.2-5.5 (1); 5.2-5.8 (2); 4.4-6.8 (3) 4.5-4.8 (4); •5 (5) 5.0-5.5 (6) •5 (7) 5.5-6.0 (4) 5.0 (4); 5.0-5.5 (8); 4.5-5.5 (9) 5.0-5.5 (4) 4.5-4.7 (1); (4.7 (2); 4.4 (3) (4.4 (3)

4.7-5.2 (2); 4.2-5.0 (3)

4.8-5.2 (3) •5.7 (3) 4.5-4.7 (1) 5.7-6.0 (3) 5.3-5.7 (7)

4.7-5.2 (1,2); 4.2-5.0 (3)

4.5-5.2 (1,2); 4.6-5.0 (3)

5.2-5.5 (1)

5.5-6.0(1); 5.2 5.8 (2)

5.5-6.0 (1);)5.5 (2); •5.8 (10); 4.4-5.0 (3) 4.4-5.2 (3) 4.7-5.2 (1,2); 4.2-5.0 (3) 4.7-5.2 (1); (4.2 (3) (4.2 (3)

5.0-5.9 (3) 4.8-5.9 (3) 5.5-6.0 (1); 5.2-5.8 (2) 4.5-4.8 (1); (4.7 (2); 4.2-4.4 (3) 5.0-5.5 (11)

a References: (1) Beamish 1976; (2) Beamish et al. 1975; (3) Harvey 1980; (4) Grande et al. 1978; (5) Schofield 1976; (6) Anonymous 1978; (7) Ahner et al. 1974; (8) Jensen and Snekvik 1972; (9) Wright and Snekvik 1978; (10) Pfeiffer and Festa 1980; (11) Runn et al. 1977.

ported that reproduction was impaired in bluntnose minnows exposed to 18/xg/liter copper. These responses also may occur in acidified lakes and streams.

Low pH also disturbs the normal action of choriolytic enzymes in embryos of the Euro- pean perch. At pH 5, normal degradation of the chorion did not occur, and hatching failure

resulted (Runn et al. 1977). The same response has been reported for Atlantic salmon (Peter- son et al. 1980a, 1980b). Also, movement of Atlantic salmon embryos within the egg, important in rupturing the outer egg membrane, were reduced at low pH (Peterson et al. 1980b).

Beamish (1976) observed that female white suckers from acidified lakes did not release

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T^BLE 4.--Values of pH found in laboratory experiments to cau•'e various' adverse effects on fish species. Duration of exposure ranged from 4 days to full life cycles. References are in parentheses. a

Increased mortality

Family and Juveniles Reduced species Embryo Fry or adults growth Other effects

Sahnonidae

Brook trout 6.5 (1) 4.4 (2) 4.5 (1) 6.5 (1) Reducedegg viability: 5.0(1) 5.6 (4) 4.5 (5) 4.1 (6) 4.6 (7) Tissue damage: 5.2(3) 4.5 (5) 6.1 (1) 3.5 (8)

Arctic char 4.8 (9)

Rainbow trout 5.5 (10) 4.3 (11) 3.6•..1 (10) 4.8 (9)

Brown trout 4.0 (12) 5.0 (5) 4.1 (5)

Atlantic sahnon 3.4-4.4 (11) 4.0 (13) 3.6 (i3) 4.3 (15) 3.9 (i5) 4.3(11) 4.0 (i2) 5.o (5)

4.0-5.5 (16) 4.1 (5)

Esocidae

Northern pike 5.0 (17)

Cyprinidae Roach 5.6 (18)

Fathead minnow 5.9 (19) 5.9 (19)

Catostomidae

White sucker 4.5 (20) 5.3 (20) 4.0 (4)

Percidae

European perch 5.6 (18) 5.5 (22)

2.1 (19) 4.5 (19)

4.5(21)

Tissue damage: 5.0 (14)

Reduced egg viability: 6.6 (19)

Ceased feeding: 4.5 (21) Bone deformity: 4.2 (21); 5.0 (4)

a References: (1) Menendez 1976; (2) Schofield and Trojnar 1980; (3) Daye and Garside 1976; (4) Trojnar 1977b; (5) Johansson et al. 1977; (6) Robinson et al. 1976; (7) Leivestad et al. 1976; (8) Daye and Garside 1975; (9) Edwards and Hjeldnes 1977; (10) Kwain 1975; (11) Daye 1980; (12) Carrick 1979; (13) Daye and Garside 1977; (14) Daye and Garside 1980; (15) Daye and Garside 1979; (16) Peterson et al. 1980a; (17) Johansson and Kihlstrom 1975; (18) Johansson and Milbrink 1976; (19) Mount 1973; (20) Trojnar 1977a; (21) Beamish 1972• (22) Runn et al. 1977.

eggs, and the species then shortly disappeared from the lake. The lack of egg deposition was coincident with an inability of females to maintain normal serum calcium levels (Beamish et al. 1975; Beamish 1976; Lockhart and Lutz 1976), which was believed to be caused by acid stress. Serum calcium, in the form of calcium phosphoproteinate, normally increases in females during the period of ovary development. Serum calcium in females exceeded that of

males by a factor of 1.4 or more in nonacidified lakes, but was less than this in acidified lakes where the species was believed to be failing (Lockhart and Lutz 1976). Ruby et al. (1977) found that depressed pH reduced the ability of flagfish oocytes to form mature eggs. Sper- matogenesis in males also was reduced by low

pH, but not to the degree that egg production was affected (Ruby et al. 1978).

Low pH may prevent spawning in some species, or restrict spawning to unsuitable areas. Johnson and Webster (1977) found that brook trout avoided water of pH less than 5 when selecting spawning sites. Agreement is generally good between the pH at which reproductive failure may occur, as shown by laboratory tests (Table 3), and the field pH known to affect a species (Table 4). However, the lowest field pH is generally higher than the lowest laboratory pH, possibly because other toxic substances, such as aluminum, may be present in field situations. Leivestad et al. (1980) and Muniz and Leivestad (1980b) have shown that water from acidified streams in Norway is more

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toxic to fish than tap water acidified to the same pH. They believe that aluminum causes the additional toxicity of stream water.

Growth

Growth of fish may either increase or decrease after acidification. Ryan and Harvey (1977) found that growth of rock bass was greater than expected in acidified lakes, but although growth of yellow perch increased at ages 1 to 3, it decreased among older fish (Ryan and Harvey 1980). Beamish (1974a) found that growth of white suckers was reduced in acidified lakes, even though invertebrate food organisms were abundant. Several laboratory studies have also shown reduced growth by several species of fish exposed to sublethal levels of pH (Table 4) but not by brown trout (Ed- wards and Hjeldnes 1977; Jacobsen 1977; Gjedrem 1980a). Also, Mudge et al. (1977) found that exposure to low pH reduced ribo- nucleic acid synthesis in brook trout, which indicates reduced protein synthesis and thus growth. Growth may increase because elimi- x•ation of competing species increases availabil- ity of food. Conversely, growth may decrease because sublethal acid stress increases metabo-

lism. While exposure to lethal pH reduces oxygen consumption, exposure to sublethal pH and aluminum increases oxygen consumption and metabolism in brown trout, and reduces growth (Rosseland 1980).

Skeletal Deformity

Skeletal deformities have been reported in white suckers from acidified lakes (Beamish et al. 1975). The same effect was produced in laboratory exposures of young white suckers to low pH (Beamish 1972; Trojnar 1977a). These deformities may be a result of decalcification of the skeleton to maintain serum osmotic concen-

tration in the face of acidosis. This response has not been observed in other species of fish.

Metal Uptake

Fish from acidified lakes may contain elevated concentrations of mercury and other metals. In Sweden, mercury contents of fish from remote lakes were negatively correlated with pH (Jernelov et al. 1975). Northern pike contained as much as 4.8 mg/kg (wet weight) mercury. Armstrong and Sloan (1980) found that the mercury content of fish from four Adirondack

lakes with low alkalinity ((15 mg/liter as CaCOa) increased between 1970 and 1978. Bloomfield et al. (1980) reported high mercury concentrations in fish in Cranberry Lake, New York. Although this lake has a summer pH of 6, it receives acidic precipitation and the pH declines to 4.5 at spring snowmelt.

Thompson et al. (1980) reported that fish from remote lakes in Ontario and southeastern

New Brunswick, Canada, have elevated levels

of mercury, apparently associated with declin- ing pH. In Muskoka Lake the average mercury content of "trout" (presumably brook trout) increased from 0.5 to 3.0 mg/kg in 4 years. Schei- der et al. (1979) showed that mercury concentrations in walleye were higher in Ontario lakes with alkalinity of 0-15 mg/liter (as CaCOs) than in lakes with higher alkalinity. Northern pike and walleyes from low-alkalinity lakes in the Boundary Waters Canoe Area in northern Minnesota also contained elevated mercury levels (Glass and Loucks 1980). Because acidification reduces alkalinity, this phenomenon may occur in acidified lakes as well. However, Suns et al. (1980) found no such effect in yellow perch from Ontario lakes; mercury content of yearling fish was not correlated with alkalinity, but it was correlated with hydrogen ion concentration.

The reasons for the observed increase in

mercury are not clear, but a number of mechanisms are possible. The transport of mercury to remote lakes by precipitation may contribute significant amounts of mercury to the system (Jernelov et al. 1975). Although mercury in the atmosphere may result largely from natural sources, acidic precipitation scavenges mercury from the air more effectively than nonacidic precipitation (Tomlinson et al. 1980). Bloom- field et al. (1980) found that more mercury was transported to Cranberry Lake from the watershed than was deposited in precipitation. The rate of methylation of inorganic mercury by microorganisms is pH-dependent, the maxi- mum occurring at pH 6; methylation is higher over the pH range 5-7 than above pH 7 (Tom- linson et al. 1980; Wood 1980). Thus at lower pH one would expect more methylmercury (the form most rapidly taken up by fish) to be present. Under aerobic conditions, mercury, and other metals such as selenium and zinc, are

more soluble at reduced pH, and are thus more available for methylation reactions (Wood

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1980). Jackson et al. (1980) reported that the normal removal of mercury and other metals from the water column to the sediments was

reduced when lake enclosures were experimentally acidified, resulting in increased concentrations of metals in the water column. Miller and

Akagi (1979) found that methylation of mercury was not affected by pH over the range of 5.0 to 7.5, but that the distribution of methylmercury between sediments and water did change, more methylmercury being found in the water at lower pH. Fish also accumulate increasing amounts of inorganic mercury as pH is reduced (Drummond et al. 1974; Tsai et al. 1975). Because only a small fraction of the mercury in an aquatic ecosystem is in fish, it is likely that the observed increases are caused by some change in biological processes in acidified lakes, rather than simply from an increased influx of mercury (Holden and Spencer 1979). Jernelov (1980) and Tomlinson et al. (1980) believe that lower pH leads to higher mercury content of fish through two major mechanisms. Firstly, greater acidity results in increased retention in the water column of mercury that otherwise would be deposited in the sediments. Fish thus are exposed to higher concentrations of dissolved mercury, which they can take up directly through their gills. Secondly, the biomass and growth rates of invertebrates and fish may decline in an acidic lake. The available mercury then will be concentrated in a smaller biomass, resulting in higher body burdens in both fish and their food. If growth rate is reduced, fish in an acidic lake will be older than fish of an

equivalent size in a nonacidic lake, and will have been accumulating mercury longer.

Other metals may behave in a manner similar to mercury. For example, the blood lead concentration in rainbow trout increased as water

pH was reduced from 10 to 6 (Hodson et al. 1978). Merlini and Pozzi (1977) found that pumpkinseeds accumulated three times more lead at pH 6 than at 7.5.

Ecosystem Effects In aquatic ecosystems, all trophic levels con-

tain organisms that are sensitive to acidification. As a result the number and diversity of species decline. However, primary production may continue at near-historical levels, and the input of allochthonous material probably is not reduced. While no measurements of secondary

production have been made, the total number of organisms and total biomass may not decline because acid-tolerant species are able to use available resources. The limiting factor for production may be a primary nutrient, such as phosphorus, rather than toxic effects of pH or metals. The disappearance of fish may result in profound changes in plant and invertebrate communities because of the replacement of vertebrate predators by invertebrate predators. Fish generally are highly size-selective in their choice of prey items, and the presence of fish affects invertebrate size distribution, diversity, and even number of species (Hall et al. 1970). In acidified lakes, the dominant fish species may function as "keystone predators." As such they would influence the nature and extent of interactions among a whole subweb of communities, and thereby the diversity and stability of the ecosystem as a whole. The disappearance of fish from an acidified lake may be of more importance to the ecosystem than the direct effects of pH and toxic metals on individual species or groups of organisms.

Vulnerability of Surface Waters to Acidification

The extent of the geographical areas that receive acidic precipitation and contain surface waters vulnerable to damage from this precipitation is unknown. A number of precipitation- chemistry networks are, or have been, in operation throughout the world, but primarily in Europe and North America (Whelpdale 1979). In North America, coverage still is very sparse and time-series data are lacking. Acidification of surface waters by precipitation has been demonstrated for only a few geographic areas. Predictions of areas of the world that may contain sensitive waters are based on bedrock ge- ology, soil chemistry, or water chemistry (Gal- loway and Cowling 1978; Altshuller and McBean 1980; Hendrey et al. 1980; McFee 1980; Root et al. 1980).

Surface waters with alkalinity of 10 mg/liter as CaCOs or less are classified as highly sensitive to acidification, and those with 10-20 mg/liter as moderately sensitive (Altshuller and McBean 1979). Brooks and Deevey (1966) reported that a high proportion of New England waters are "extremely soft," having less than 10 mg/liter bicarbonate. A large portion of eastern Canada also contains surface waters with alkalinity less

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than 10 mg/liter as CaCOa (Altshuller and McBean 1979), as do most lakes in the Bound- ary Waters Canoe Area in northern Minnesota (Glass and Loucks 1980). This region of soft surface waters extends into Wisconsin and

northern Michigan. Other indices of sensitivity have been pro-

posed. The calcite saturation index (CSI) proposed by Kramer (1976) and Galloway et al. (1978) is a measure of neutralizing capacity based on alkalinity, calcium, and pH. This index may be more useful than alkalinity alone in predicting whether or not a lake will become acidified, and especially in predicting effects of acidity on fish (calcium is known to mediate physiological effects of acid and metals). On the basis of the CSI index, Glass and Loucks (1980) concluded that 7% of the lakes in the Boundary Waters Canoe Area were highly susceptible to acidification, 30% were susceptible, and an additional 34% probably were susceptible. A model for identifying and measuring acidification of surface waters for which historical data are

lacking, proposed by Henriksen (1979), was derived from an empirical fit of pH and calcium data for a large number of lakes. The model correctly predicted pH of surface waters in several different acidified-lake districts

(Wright et al. 1980), and predicted loss of fish from acidified lakes in New York (Hendrey et al. 1980).

Almer et al. (1978) determined that important pH changes would occur in very sensitive lakes at a sustained sulfate loading of 15 kg SOd hectare per year (0.5 g S/m 2 per year). Changes would occur in less sensitive lakes at sustained

sulfate loadings above 30 kg SOdhectare per year (1.0 g S/m 2 per year). These relationships have not been tested in North America, but substantial portions of the northeastern United States and eastern Canada receive precipitation-sulfate loadings of 10-15 kg SOdhectare per year, and loadings are as high as 36 kg SOd hectare per year in some areas (National At- mospheric Deposition Program 1980).

Surveys of the abundance and distribution of acidified lakes have been made in Scandinavia

and parts of North America, and the probabil- ity of future acidification has been estimated. Of the 90,000 lakes in Sweden, about 20,000, located in an area of 100,000 km •, are now acidic (Bengtsson et al. 1980). It is predicted that half of the lakes will be acidic within 10 years (Dickson 1975, 1980). Most of the more than

200,000 lakes in Norway are sensitive to acidification; in the southernmost lakes, pH usually is 5 or less. Fish populations have been adversely affected in 20% of the southern lakes (Lei- vestad et al. 1976; Overrein 1976). Adverse effects on fish populations occur in an area of 33,000 km •, and fish are virtually extinct in an area of 13,000 km • (Sevaldrud et al. 1980). In Ontario, at least 200 lakes are known to be ad-

versely affected by acidic precipitation, and 48,000 others are expected to follow suit over the next 20 years (Parrott 1979; Harvey 1980).

The accuracy with which acidification can be predicted by the CSI or other indices is open to question. The response of a given body of water to a specific input of acid is governed by a complex mixture of factors, and estimates of biologic response based on chemical buffering capacity alone may not be sufficient. For example, Schindler• et al. (1980) found that the pH of Lake 223 in the Experimental Lakes Area, Ontario decreased less than calculated when acid was added. Sulfate reduction by bacteria increased as additions of sulfuric acid in-

creased. Iron mobilized from anoxic sediments

caused bicarbonate from bacterial respiration to accumulate; this balanced the iron charge, and increased the acid-buffering capacity. It is not known how widespread or predictable this reaction is. Also, acidic precipitation contains a significant nitric acid component that is increasing, and that may enter into a different set of reactions. Predictions of future impacts or trends based on current knowledge should be regarded with caution. There is an urgent need for information about the factors that control

the vulnerability of surface waters to acidification.

Remedial Action

The obvious solution to the "acid rain" problem is a reduction in anthropogenic emissions of sulfur and nitrogen oxides. If all carbonate minerals have not been leached from a wa-

tershed, surface waters should gradually increase in pH as the input of acid decreases. However, if all carbonates have been leached by acid, surface water pH probably will reflect the pH of precipitation, unless basic chemicals are added to the watershed. Political and socio-

logical realities dictate that reduction of sulfur and nitrogen oxides will not be forthcoming in the near future, and will never approach zero. The current lack of dose-response information

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makes it impossible to estimate the improvement that would result from specific reductions in these emissions.

The economic losses caused by acidic precipitation should be weighed against the cost of emission reductions, but such costs currently defy accurate estimation. Commercially fished populations of Atlantic salmon have been lost in Sweden and Norway, and elimination of sport fisheries for trout by acidification has cost New York an estimated annual loss of $1 million in tourism (Altshuller and McBean 1979). Tuomi (1981) estimated that the economic value of sport fisheries in Canada is equivalent to that of commercial fisheries. Loss of sport fisheries would result in severe economic loss in

eastern Canada. None of these costs take into

account the intangible benefits of a healthy environment.

Technology for control of emissions of acidic gases is available (Loucks 1981 a). Likens (1976) estimated that an expenditure of more than $4 billion would be required to reduce the emission of SO2 by half in the United States. This cost possibly could be reduced by application of recovered sulfur as a fertilizer in sul-

fur-impoverished areas, such as the southeastern United States. Duckenfield (1976) estimated that a 70% reduction in sulfur oxide emissions

in western Europe (from 22 million to 6-8 million tonnes per year) would cost $5 billion per year for 10 years. Barnes (1979) calculated the cost of a 60% reduction in western Europe at $10 billion per year. Persson (1976) estimated that a 60% reduction for all of Europe (from 60 million to 25 million tonnes per year) would cost about $9 billion per year. Loucks (1981b) estimated that the cost of abatement of SO2

emissions is approximately equal to the dollar value of damage caused by acid rain, not including benefits to human health, forest effects, or esthetics.

If the source of the problem cannot be treated, it is possible to treat the symptoms in acidified lakes and streams. Such remedial mea-

sures include chemical neutralization of acids,

selective breeding of acid-tolerant fish, and stocking of hatchery fish.

Neutralization

Addition of chemicals is the most widely practiced method to date of neutralizing acids. The least expensive and most widely available neutralizing agent is ground limestone (CaCO3

or a mixture of CaCOa and MgCOa). In Swe- den, 900 lakes and rivers have been limed in an area of 6,000 km 2 (Dickson 1978; Bengtsson et al. 1980). Direct application of lime to the water gave the best pH control at the lowest cost. Ap- plication rates of 10-20 g/m a, or 50-75 kg/hectare, are expected to yield acceptable pH for 5 years. If applied to the watershed, rather than directly to the water, doses must be 100 times higher. The major problem is that of keeping the pH high during periods of snowmelt and other high flow. The River Hogvadsan has required lime applications totalling 10,700 tonnes over 4 years in the 476-km 2 watershed, at a cost of more than $500,000. The river still experi- ences pH declines at spring runoff (Edman and Fleischer 1980).

The biological effects of liming have been fa- vorable. The decomposition rate of organic matter increases and Sphagnum mats are reduced. Species diversity of phytoplankton returns to normal within 1-2 years. Zooplankton increases in abundance, but acid-sensitive species generally require 4 years or longer to reappear. The benthic invertebrate community returns to pre-acid conditions in time, if im- migration from nearby waters is possible. Re- production of European perch, roach, and At- lantic salmon has been restored in acidic waters

that have been limed (Dickson 1978; Bengtsson et al. 1980; Muniz and Leivestad 1980a).

Intensive studies of the effects of neutral-

ization on lake chemistry and biota have been carried out in Ontario. A combination of

Ca(OH)2 and CaCOa was superior to either chemical alone in quickly returning pH to neutral and establishing a more stable buffer system (Scheider and Dillon 1976). The addition of a base to the lakes also lowered heavy metal concentrations (by 13-83%, depending on metal and lake) but did not affect other major ions (Yan et al. 1977; Dillon et al. 1979). A lake of 21 hectares with a mean depth of 9.4 m required 20,000 kg of Ca(OH)2 and 14,000 kg of CaCOa to raise the pH from 4.5 to 7.0 (Scheider et al. 1975). If only CaCOa were used, the quan- tities required to raise pH would be much greater because its reactivity is low and insolu- ble reaction products are precipitated.

The Ontario biota responded variably to neutralization. When pH was altered rapidly (raised three units within a few weeks), the standing stock of phytoplankton, zooplankton, and benthic invertebrates declined initially.

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694

The bacteria were the first organisms to respond positively, the heterotrophic bacteria increasing by several orders of magnitude and replacing aciduric bacteria as the dominant organisms (Scheider and Dillon 1976). The year after neutralization, phytoplankton biomass returned to pretreatment levels, and species composition began to return to a more normal distribution, with the proportion of Chrysophyta increasing. Similar changes were observed in tube experiments (Yah and Stokes 1978). Phy- toplankton biomass was increased above pretreatment levels only by the addition of phosphorus (Dillon et al. 1979). Zooplankton returned to pretreatment abundance within 3 years, but species composition was unaltered from that before treatment. The addition of

phosphorus increased zooplankton standing stock, but species composition did not change (Dillon et al. 1979). Benthic invertebrates began to recover toward pretreatment abundance only slowly, and did not respond to phosphorus additions; species composition was unaltered (Scheider et al. 1976).

Fish stocked in several lakes in Ontario fol-

lowing neutralization did not survive. Rainbow trout suspended in cages in two of these lakes lived only a few days, whereas they survived at least 50 days in a nonacid lake. Despite reductions in heavy metals after neutralization, copper and possibly also zinc and nickel concentrations remained high enough to have caused the fish mortality (Yah et al. 1979). Similar problems have been experienced in some Swedish waters. There is a transition period after lime has been added when pH has increased but metals (aluminum in Sweden) have not precipitated, and fish mortality may occur (Bengtsson et al. 1980).

After lime was applied to acidic lakes totalling 332 hectares in the Adirondack Mountains of

New York in 1979, fishable populations of brook trout were maintained (M. Pfeiffer, New York Department of Environmental Conser- vation, personal communication). Biological effects on aquatic organisms other than fish have not been assessed.

In Sweden, $12 million was scheduled to be spent in 1977-1980 for application of lime. In 1979 the cost of applying lime was $50-70 per ton, or about $50-60 per surface hectare of water treated (Bengtsson et al. 1980). Ducken- field (1976) estimated the cost as $100-125 per

hectare. In New York, liming costs ranged from $55 to $470 per hectare and averaged $150, and the estimated annual cost of liming all acidic lakes in New York is $5 million (Horn et al. 1980). Barnes (1979) estimated that all acidic waters in Scandinavia could be neutralized for

$150 million per year--a price that compares favorably with the annual cost of reducing sulfur oxide emissions at the source. Thus if acid-

ification of surface waters is the only adverse effect of acid rain, chemical neutralization is an economically feasible remedial technique.

Selective Breeding

Attempts to acclimate fish to low pH generally have failed (Lloyd and Jordan 1964; Falk and Dunson 1977; Swarts et al. 1978; Daye 1980). However, wild fish from acidic waters are more tolerant of acidity than hatchery fish (Dunson and Martin 1973; McWilliams 1980). Different genetic strains of brook and brown trout differ in tolerance to acid, and this trait is heritable (Gjedrem 1976, 1980b; Robinson et al. 1976; Edwards and Gjedrem 1979). Work to develop acid-tolerant strains for introduction into acidic lakes is under way in New York and Norway (C. Schofield, Cornell University, personal communication). Hybrid brook trout (produced by crossing selected Canadian strains with a New York domestic strain) stocked in acidic lakes in New York have shown

excellent survival and growth (M. Pfeiffer, New York Department of Environmental Conser- vation, personal communication).

Different strains of fish of the same species spawn at different times. A late-spawning strain could avoid the stress associated with episodes of acidity such as fall rains or spring snowmelt (Potts and Brown 1979).

Although acid-tolerant fish appear to thrive on food organisms that also tolerate these conditions, many valuable food organisms are lost in acidified lakes, and fish production might be reduced even if a genetic strain could be produced that survived the acid and metal concen-

trations. Genetic selection appears to hold promise for establishing or maintaining recreational trout fishing in lakes in the early stages of acidification. Fish production and distribution facilities already exist that could be adapt- ed to this procedure. However, if atmospheric emissions continue unabated or increase, the

acidification process may further depress pH to

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a point where the stocking of genetically selected fish becomes ineffective as a management measure.

Maintenance Stocking

If the loss of fish populations from acidic lakes is primarily from failure to reproduce, the introduction of hatchery fish could be used to maintain fishable populations. In some cases, hatchery fish have died soon after they were stocked in acidic lakes, especially if the fish came from hard-water hatcheries (Schofield 1965; C. Thoits, New Hampshire Department of Fish and Game, personal communication). However, I found populations of hatchery-produced brook trout in New Hampshire, Maine, and Vermont lakes with pH below 5, whereas similar lakes that were not stocked were devoid

of fish (unpublished data).

Other Actions

Other techniques for remedial action have been reported or suggested, but not yet tested on a large scale. Rather than neutralization of entire lakes, the treatment of embayments may be sufficient to provide refuges from acidic conditions during spring. Fish are able to locate tributary streams of high pH in an otherwise acidic lake (C. Schofield, Cornell University, personal communication). Gunn and Keller (1980) reported that in situ incubation of eyed rainbow trout eggs in spawning boxes filled with limestone chips was a promising technique for maintaining fish populations in acidified lakes. Because metals are a major cause of fish mortality in acidified lakes, treatment with metal-chelating agents may be a successful management measure. Gorham (1981) suggested the possibility of breeding acid-tolerant strains of sulfate and nitrate reducing bacteria that would reduce acidity in lakes receiving acidic precipitation.

Summary

Precipitation in certain areas of the globe has become highly acidic in the last few decades, and it has had a profound effect on surface waters with low neutralizing capacity. Such waters have decreased alkalinity and pH, and increased metals and organic compounds. De- composers, algae, macrophytes, invertebrates, and fish have been affected. These effects in-

clude reduced abundance, production, or

growth, and losses of sensitive species; these losses result in altered species compositions of the various trophic groups, and in reduced species diversity. The magnitude and distribution of affected areas is still largely unknown, and projections for future effects are difficult to make. The master environmental variable

appears to be hydrogen ion, but the effects on a particular species may result from acid stress, increased concentrations of metals (leached from soil or sediment by acid or released during fossil fuel combustion), or synergistic inter- action between acid and metals. There may be secondary effects resulting from changes in food or predator organisms. Toxic organic compounds are associated with acidic precipitation but their significance is poorly understood.

The importance of acidic precipitation to us lies in the loss of recreational and commercial

fishing, with associated intrinsic and economic values; in the potential contamination of edible fish with toxic metals and organic compounds that render them unsafe to eat; and in the potential contamination of drinking water with toxic metals or organic compounds.

The ultimate solution to the problem is a reduction in anthropogenic emissions of sulfur and nitrogen oxides. However, the degree of reduction required to produce a given degree of improvement in sensitive environments is unknown. Possible interim remedial or mitigative actions include reduction in acidity of selected environments by addition of basic ma- terials, and hatchery production of fish that can survive in the acidified waters.

Acknowledgments I am indebted to numerous scientists who

generously supplied me with preprints and re- prints of their research findings so that I could include them in this review. S. A. Norton, C. L.

Schofield, J. G. Wiener, and P. C. Baumann supplied helpful critiques of early versions of the manuscript. J. H. Mowatt maintained the extensive reprint file and typed the manuscript.

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Documents

Acidic Precipitation and Its Consequences for Aquatic Ecosystems: A Review